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United States Office of Research and EPA/600/R-98/128 

Environmental Protection Development September 1998 

Agency Washington, DC 20460 

■SEPA Technical Protocol for 

Evaluating Natural 
Attenuation of Chlorinated 
Solvents in Ground Water 













































TECHNICAL PROTOCOL FOR EVALUATING NATURAL 
ATTENUATION OF CHLORINATED SOLVENTS IN 

GROUND WATER 


by 

Todd H. Wiedemeier 
Parsons Engineering Science, Inc. 

Pasadena, California 

Matthew A. Swanson, David E. Moutoux, and E. Kinzie Gordon 
Parsons Engineering Science, Inc. 

Denver, Colorado 

John T. Wilson, Barbara H. Wilson, and Donald H. Kampbell 
United States Environmental Protection Agency 
National Risk Management Research Laboratory 
Subsurface Protection and Remediation Division 
Ada, Oklahoma 

Patrick E. Haas, Ross N. Miller and Jerry E. Hansen 
Air Force Center for Environmental Excellence 
Technology Transfer Division 
Brooks Air Force Base, Texas 

Francis H. Chapelle 
United States Geological Survey 
Columbia, South Carolina 


IAG #RW57936164 

Project Officer 
John T. Wilson 

National Risk Management Research Laboratory 
Subsurface Protection and Remediation Division 

Ada, Oklahoma 


NATIONAL RISK MANAGEMENT RESEARCH LABORATORY 
OFFICE OF RESEARCH AND DEVELOPMENT 
U. S. ENVIRONMENTAL PROTECTION AGENCY 
CINCINNATI, OHIO 45268 


NOTICE 


The information in this document was developed through a collaboration between the U.S. 
EPA (Subsurface Protection and Remediation Division, National Risk Management Research 
Laboratory, Robert S. Kerr Environmental Research Center, Ada, Oklahoma [SPRD]) and the U.S. 
Air Force (U.S. Air Force Center for Environmental Excellence, Brooks Air Force Base, Texas 
[AFCEE]). EPA staff were primarily responsible for development of the conceptual framework for 
the approach presented in this document; staff of the U.S. Air Force and their contractors also 
provided substantive input. The U.S. Air Force was primarily responsible for field testing the 
approach presented in this document. Through a contract with Parsons Engineering Science, Inc., 
the U.S. Air Force applied the approach at chlorinated solvent plumes at a number of U.S. Air 
Force Bases. EPA staff conducted field sampling and analysis with support from ManTech 
Environmental Research Services Corp., the in-house analytical support contractor for SPRD. 

All data generated by EPA staff or by ManTech Environmental Research Services Corp. were 
collected following procedures described in the field sampling Quality Assurance Plan for an in- 
house research project on natural attenuation, and the analytical Quality Assurance Plan for ManTech 
Environmental Research Services Corp. 

This protocol has undergone extensive external and internal peer and administrative review by 
the U.S. EPA and the U.S. Air Force. This EPA Report provides technical recommendations, not 
policy guidance. It is not issued as an EPA Directive, and the recommendations of this EPA Report 
are not binding on enforcement actions carried out by the U.S. EPA or by the individual States of 
the United States of America. Neither the United States Government (U.S. EPA or U.S. Air Force), 
Parsons Engineering Science, Inc., or any of the authors or reviewers accept any liability or 
responsibility resulting from the use of this document. Implementation of the recommendations of 
the document, and the interpretation of the results provided through that implementation, are the 
sole responsibility of the user. 

Mention of trade names or commercial products does not constitute endorsement or 
recommendation for use. 


TDh/jy 

. Oy A/i 

nn 


tiGCb 3 .' 12 '?? 

17 2 - Joi 


11 



FOREWORD 


The U.S. Environmental Protection Agency is charged by Congress with protecting the Nation’s 
land, air, and water resources. Under a mandate of national environmental laws, the Agency strives 
to formulate and implement actions leading to a compatible balance between human activities and 
the ability of natural systems to support and nurture life. To meet these mandates, EPA’s research 
program is providing data and technical support for solving environmental problems today and 
building a science knowledge base necessary to manage our ecological resources wisely, understand 
how pollutants affect our health, and prevent or reduce environmental risks in the future. 

The National Risk Management Research Laboratory is the Agency’s center for investigation 
of technological and tnanagement approaches for reducing risks from threats to human health and 
the environment. The focus of the Laboratory’s research program is on methods for the prevention 
and control of pollution to air, land, water, and subsurface resources; protection of water quality in 
public water systems; remediation of contaminated sites and ground water; and prevention and 
control of indoor air pollution. The goal of this research effort is to catalyze development and 
implementation of innovative, cost-effective environmental technologies; develop scientific and 
engineering information needed by EPA to support regulatory and policy decisions; and provide 
technical support and information transfer to ensure effective implementation of environmental 
regulations and strategies. 

The site characterization processes applied in the past are frequently inadequate to allow an 
objective and robust evaluation of natural attenuation. Before natural attenuation can be used in the 
remedy for contamination of ground water by chlorinated solvents, additional information is required 
on the three-dimensional flow field of contaminated ground water in the aquifer, and on the physical, 
chemical and biological processes that attenuate concentrations of the contaminants of concern. 
This document identifies parameters that are useful in the evaluation of natural attenuation of 
chlorinated solvents, and provides recommendations to analyze and interpret the data collected 
from the site characterization process. It will also allow ground-water remediation managers to 
incorporate natural attenuation into an integrated approach to remediation that includes an active 
remedy, as appropriate, as well as natural attenuation. 


Clinton W. Hall, Director 

Subsurface Protection and Remediation Division 

National Risk Management Research Laboratory 


in 





























TABLE OF CONTENTS 


Notice.ii 

Foreword.iii 

Acknowledgments.viii 

List of Acronyms and Abbreviations.ix 

Definitions.xii 

SECTION 1 INTRODUCTION.1 

1.1 APPROPRIATE APPLICATION ON NATURAL ATTENUATION.2 

1.2 ADVANTAGES AND DISADVANTAGES.4 

1.3 LINES OF EVIDENCE.6 

1.4 SITE CHARACTERIZATION.7 

1.5 MONITORING.9 

SECTION 2 PROTOCOL FOR EVALUATING NATURAL ATTENUATION. 11 

2.1 REVIEW AVAILABLE SITE DATA AND DEVELOP PRELIMINARY 

CONCEPTUAL MODEL.13 

2.2 INITIAL SITE SCREENING.15 

2.2.1 Overview of Chlorinated Aliphatic Hydrocarbon Biodegradation.15 

2.2.1.1 Mechanisms of Chlorinated Aliphatic Hydrocarbon Biodegradation.23 

2.2.1.1.1 Electron Acceptor Reactions (Reductive Dehalogenation).23 

2.2.1.1.2 Electron Donor Reactions.25 

2.2.1.1.3 Cometabolism.25 

2.2.1.2 Behavior of Chlorinated Solvent Plumes.26 

2.2.1.2.1 Type 1 Behavior.26 

2.2.1.2.2 Type 2 Behavior.26 

2.2.1.2.3 Type 3 Behavior.26 

2.2.1.2.4 Mixed Behavior.27 

2.2.2 Bioattenuation Screening Process.27 

2.3 COLLECT ADDITIONAL SITE CHARACTERIZATION DATA IN 

SUPPORT OF NATURAL ATTENUATION AS REQUIRED.34 

2.3.1 Characterization of Soils and Aquifer Matrix Materials.37 

2.3.2 Ground-water Characterization.38 

2.3.2.1 Volatile and Semivolatile Organic Compounds.38 

2.3.2.2 Dissolved Oxygen.38 

2.3.2.3 Nitrate.39 

2.3.2.4 Iron (II).39 

2.3.2.5 Sulfate.39 

2.3.2.6 Methane.39 

23 . 2.1 Alkalinity.39 

2.3.2.8 Oxidation-Reduction Potential.40 

2.3.2.9 Dissolved Hydrogen.40 

2.3.2.10 pH, Temperature, and Conductivity.41 

2.3.2.11 Chloride.42 

2.3.3 Aquifer Parameter Estimation.42 

2.3.3.1 Hydraulic Conductivity.42 

2.3.3.1.1 Pumping Tests in Wells.43 

2.3.3.1.2 Slug Tests in Wells.43 

2.3.3.1.3 Downhole Flowmeter.43 


v 














































2.3.3.2 Hydraulic Gradient.44 

2.3.3.3 Processes Causing an Apparent Reduction in 

Total Contaminant Mass.44 

2.3.4 Optional Confirmation of Biological Activity.45 

2.4 REFINE CONCEPTUAL MODEL, COMPLETE PRE-MODELING 

CALCULATIONS, AND DOCUMENT INDICATORS OF NATURAL 
ATTENUATION.45 

2.4.1 Conceptual Model Refinement.46 

2.4.1.1 Geologic Logs.46 

2.4.1.2 Cone Penetrometer Logs.46 

2.4.1.3 Hydrogeologic Sections.46 

2.4.1.4 Potentiometric Surface or Water Table Map(s).47 

2.4.1.5 Contaminant and Daughter Product Contour Maps.47 

2.4.1.6 Electron Acceptor, Metabolic By-product, and 

Alkalinity Contour Maps.47 

2.4.2 Pre-Modeling Calculations.48 

2.4.2.1 Analysis of Contaminant, Daughter Product, Electron Acceptor, 

Metabolic By-product, and Total Alkalinity Data.48 

2.4.2.2 Sorption and Retardation Calculations.49 

2.4.2.3 NAPL/Water Partitioning Calculations.49 

2.4.2.4 Ground-water Flow Velocity Calculations.49 

2.4.2.5 Biodegradation Rate-Constant Calculations.49 

2.5 SIMULATE NATURAL ATTENUATION USING SOLUTE FATE AND 

TRANSPORT MODELS.49 

2.6 CONDUCT A RECEPTOR EXPOSURE PATHWAYS ANALYSIS.50 

2.7 EVALUATE SUPPLEMENTAL SOURCE REMOVAL OPTIONS.50 

2.8 PREPARE LONG-TERM MONITORING PLAN.50 

2.9 PRESENT FINDINGS.52 

SECTION 3 REFERENCES.53 

APPENDIX A.A1 -1 

APPENDIX B. Bl-1 

APPENDIX C. Cl-1 


vi 




























FIGURES 


No. Title Page 

2.1 Natural attenuation of chlorinated solvents flow chart.12 

2.2 Reductive dehalogenation of chlorinated ethenes.24 

2.3 Initial screening process flow chart.28 

2.4 General areas for collection of screening data.31 

2.5 A cross section through a hypothetical release.36 

2.6 A stacked plan representation of the plumes that may develop from the 

hypothetical release.36 

2.7 Hypothetical long-term monitoring strategy.51 


TABLES 

No. Title Page 

i. Contaminants with Federal Regulatory Standards.xiv 

2.1 Soil, Soil Gas, and Ground-water Analytical Protocol.16 

2.2 Objectives for Sensitivity and Precision to 

Implement the Natural Attenuation Protocol.21 

2.3 Analytical Parameters and Weighting for Preliminary Screening for 

Anaerobic Biodegradation Processes.29 

2.4 Interpretation of Points Awarded During Screening Step 1.32 

2.5 Range of Hydrogen Concentrations for a Given Terminal 

Electron-Accepting Process.41 


vii 















ACKNOWLEDGMENTS 


The authors would like to thank Dr. Robert Hinchee, Doug Downey, and Dr. Guy Sewell for their 
contributions and their extensive and helpful reviews of this manuscript. Thanks also to Leigh 
Alvarado Benson, R. Todd Herrington, Robert Nagel, Cindy Merrill, Peter Guest, Mark Vesseley, 
John Hicks, and Saskia Hoffer for their contributions to this project. 


vm 


AAR 

AFB 

AFCEE 

ASTM 

bgs 

BRA 

BRAC 

BTEX 

CAP 

CERCLA 

cfm 

CFR 

COPC 

CPT 

CSM 

DAF 

DERP 

DNAPL 

DO 

DOD 

DQO 

EE/CA 

FS 

gpd 

G 

r 

HDPE 

HSSM 

HSWA 

ID 

IDW 

IRP 

L 

LEL 

LNAPL 

LUFT 

MAP 

MCL 


LIST OF ACRONYMS AND ABBREVIATIONS 

American Association of Railroads 
Air Force Base 

Air Force Center for Environmental Excellence 
American Society for Testing and Materials 

below ground surface 
baseline risk assessment 
Base Realignment and Closure 
benzene, toluene, ethylbenzene, xylenes 

corrective action plan 

Comprehensive Environmental Response, Compensation and Liability 
Act 

cubic feet per minute 
Code of Federal Regulations 
chemical of potential concern 
cone penetrometer testing 
conceptual site model 

dilution/attenuation factor 

Defense Environmental Restoration Program 

Dense Nonaqueous Phase Liquid 

dissolved oxygen 

Department of Defense 

data quality objective 

engineering evaluation/cost analysis 

feasibility study 

gallons per day 

standard (Gibbs) free energy 

high-density polyethylene 

Hydrocarbon Spill Screening Model 

Hazardous and Solid Waste Amendments of 1984 

inside-diameter 
investigation derived waste 
Installation Restoration Program 

liter 

lower explosive limit 
light nonaqueous-phase liquid 
leaking underground fuel tank 

management action plan 
maximum contaminant level 


IX 


MDL 

method detection limit 

Pg 

microgram 

Mg/kg 

microgram per kilogram 

Pg/L 

microgram per liter 

mg 

milligram 

mg/kg 

milligrams per kilogram 

mg/L 

milligrams per liter 

mg/m 3 

milligrams per cubic meter 

mm Hg 

millimeters of mercury 

MOC 

method of characteristics 

MOGAS 

motor gasoline 

NAPL 

nonaqueous-phase liquid 

NCP 

National Contingency Plan 

NFRAP 

no further response action plan 

NOAA 

National Oceanographic and Atmospheric Administration 

NOEL 

no-observed-effect level 

NPL 

National Priorities List 

OD 

outside-diameter 

ORP 

oxidation-reduction potential 

OSHA 

Occupational Safety and Health Administration 

OSWER 

Office of Solid Waste and Emergency Response 

PAH 

polycyclic aromatic hydrocarbon 

PEL 

permissible exposure limit 

POA 

point-of-action 

POC 

point-of-compliance 

POL 

petroleum, oil, and lubricant 

ppmv 

parts per million per volume 

psi 

pounds per square inch 

PVC 

polyvinyl chloride 

QA 

quality assurance 

QC 

quality control 

RAP 

remedial action plan 

RBCA 

risk-based corrective action 

RBSL 

risk-based screening level 

redox 

reduction/oxidation 

RFI 

RCRA facility investigation 

RI 

remedial investigation 

RME 

reasonable maximum exposure 

RPM 

remedial project manager 

SAP 

sampling and analysis plan 

SARA 

Superfund Amendments and Reauthorization Act 

scfm 

standard cubic feet per minute 

SPCC 

spill prevention, control, and countermeasures 


X 


SSL 

soil screening level 

SSTL 

SVE 

SVOC 

site-specific target level 
soil vapor extraction 
semivolatile organic compound 

TC 

TCLP 

TI 

* TMB 

TOC 

TPH 

TRPH 

TVH 

TVPH 

TWA 

toxicity characteristic 

toxicity-characteristic leaching procedure 

technical impracticability 

trimethylbenzene 

total organic carbon 

total petroleum hydrocarbons 

total recoverable petroleum hydrocarbons 

total volatile hydrocarbons 

total volatile petroleum hydrocarbons 

time-weighted-average 

UCL 

US 

USGS 

UST 

upper confidence limit 

United States 

US Geological Survey 
underground storage tank 

VOCs 

volatile organic compounds 


XI 


DEFINITIONS 


Aerobe : bacteria that use oxygen as an electron acceptor. 

Anabolism : The process whereby energy is used to build organic compounds such as enzymes and 

nucleic acids that are necessary for life functions. In essence, energy is derived from catabolism, 
stored in high-energy intermediate compounds such as adenosine triphosphate (ATP), guanosine 
triphosphate (GTP) and acetyl-coenzyme A, and used in anabolic reactions that allow a cell to 
grow. 

Anaerobe : Organisms that do not require oxygen to live. 

Area of Attainment : The area over which cleanup levels will be achieved in the ground water. It 

encompasses the area outside the boundary of any waste remaining in place and up to the boundary 
of the contaminant plume. Usually, the boundary of the waste is defined by the source control 
remedy. Note: this area is independent of property boundaries or potential receptors - it is the 
plume area which the ground water must be returned to beneficial use during the implementation of 
a remedy. 

Anthropogenic: Man-made. 

Autotrophs : Microorganisms that synthesize organic materials from carbon dioxide. 

Catabolism: The process whereby energy is extracted from organic compounds by breaking them down 
into their component parts. 

Coefficient of Variation: Sample standard deviation divided by the mean. 

Cofactor: A small molecule required for the function of an enzyme. 

Cometabolism: The process in which a compound is fortuitously degraded by an enzyme or cofactor 
produced during microbial metabolism of another compound. 

Daughter Product: A compound that results directly from the biodegradation of another. For example 
cfs-l,2-dichloroethene (c/s-1,2-DCE)is commonly a daughter product of trichloroethene (TCE). 

Dehydrohalogenation: Elimination of a hydrogen ion and a halide ion resulting in the formation of an 
alkene. 

Diffusion : The process whereby molecules move from a region of higher concentration to a region of 
lower concentration as a result of Brownian motion. 

Dihaloelimination: Reductive elimination of two halide substituents resulting in formation of an alkene. 

Dispersivity: A property that quantifies mechanical dispersion in a medium. 

Effective Porosity: The percentage of void volume that contributes to percolation; roughly equivalent to 
the specific yield. 

Electron Acceptor: A compound capable of accepting electrons during oxidation-reduction reactions. 
Microorganisms obtain energy by transferring electrons from electron donors such as organic 
compounds (or sometimes reduced inorganic compounds such as sulfide) to an electron acceptor. 
Electron acceptors are compounds that are relatively oxidized and include oxygen, nitrate, 
iron (III), manganese (IV), sulfate, carbon dioxide, or in some cases the chlorinated aliphatic 
hydrocarbons such as perchloroethene (PCE), TCE, DCE, and vinyl chloride. 

Electron Donor: A compound capable of supplying (giving up) electrons during oxidation-reduction 
reactions. Microorganisms obtain energy by transferring electrons from electron donors such as 
organic compounds (or sometimes reduced inorganic compounds such as sulfide) to an electron 
acceptor. Electron donors are compounds that are relatively reduced and include fuel 
hydrocarbons and native organic carbon. 

Electrophile: A reactive species that accepts an electron pair. 

Elimination: Reaction where two groups such as chlorine and hydrogen are lost from adjacent carbon 
atoms and a double bond is formed in their place. 

Epoxidation: A reaction wherein an oxygen molecule is inserted in a carbon-carbon double bond and an 
epoxide is formed. 


Xll 


Facultative Anaerobes: microorganisms that use (and prefer) oxygen when it is available, but can also use 
alternate electron acceptors such as nitrate under anaerobic conditions when necessary. 

Fermentation : Microbial metabolism in which a particular compound is used both as an electron donor 
and an electron acceptor resulting in the production of oxidized and reduced daughter products. 

Heterotroph : Organism that uses organic carbon as an external energy source and as a carbon source. 

Hydraulic Conductivity : The relative ability of a unit cube of soil, sediment, or rock to transmit water. 

Hydraulic Head : The height above a datum plane of the surface of a column of water. In the 

groundwater environment, it is composed dominantly of elevation head and pressure head. 

Hydraulic Gradient : The maximum change in head per unit distance. 

Hydrogenolysis : A reductive reaction in which a carbon-halogen bond is broken, and hydrogen replaces 
the halogen substituent. 

Hydroxylation : Addition of a hydroxyl group to a chlorinated aliphatic hydrocarbon. 

Lithotroph : Organism that uses inorganic carbon such as carbon dioxide or bicarbonate as a carbon 
source and an external source of energy. 

Mechanical Dispersion : A physical process of mixing along a flow path in an aquifer resulting from 
differences in path length and flow velocity. This is in contrast to mixing due to diffusion. 

Metabolic Byproduct: A product of the reaction between an electron donor and an electron acceptor. 
Metabolic byproducts include volatile fatty acids, daughter products of chlorinated aliphatic 
hydrocarbons, methane, and chloride. 

Monooxygenase: A microbial enzyme that catalyzes reactions in which one atom of the oxygen molecule 
is incorporated into a product and the other atom appears in water. 

Nucleophile: A chemical reagent that reacts by forming covalent bonds with electronegative atoms and 
compounds. 

Obligate Aerobe: Microorganisms that can use only oxygen as an electron acceptor. Thus, the presence 
of molecular oxygen is a requirement for these microbes. 

Obligate Anaerobes: Microorganisms that grow only in the absence of oxygen; the presence of molecular 
oxygen either inhibits growth or kills the organism. For example, methanogens are very sensitive 
to oxygen and can live only under strictly anaerobic conditions. Sulfate reducers, on the other 
hand, can tolerate exposure to oxygen, but cannot grow in its presence (Chapelle, 1993). 

Performance Evaluation Well: A ground-water monitoring well placed to monitor the effectiveness of 
the chosen remedial action. 

Porosity: The ratio of void volume to total volume of a rock or sediment. 

Respiration: The process of coupling oxidation of organic compounds with the reduction of inorganic 
compounds, such as oxygen, nitrate, iron (III), manganese (IV), and sulfate. 

Solvolysis: A reaction in which the solvent serves as the nucleophile. 


Xlll 


Table i: Contaminants with Federal Regulatory Standards Considered in this Document 


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SECTION 1 
INTRODUCTION 


Natural attenuation processes (biodegradation, dispersion, sorption, volatilization) affect the 
fate and transport of chlorinated solvents in all hydrologic systems. When these processes are 
shown to be capable of attaining site-specific remediation objectives in a time period that is reasonable 
compared to other alternatives, they may be selected alone or in combination with other more 
active remedies as the preferred remedial alternative. Monitored Natural Attenuation (MNA) is a 
term that refers specifically to the use of natural attenuation processes as part of overall site 
remediation. The United States Environmental Protection Agency (U.S. EPA) defines monitored 
natural attenuation as (OSWER Directive 9200.4-17, 1997): 

The term “monitored natural attenuation, ” as used in this Directive, refers 
to the reliance on natural attenuation processes (within the context of a carefully 
controlled and monitored clean-up approach) to achieve site-specific remedial 
objectives within a time frame that is reasonable compared to other methods. The 
“natural attenuation processes ” that are at work in such a remediation approach 
include a variety of physical, chemical, or biological processes that, under favorable 
conditions, act without human intervention to reduce the mass, toxicity, mobility, 
volume, or concentration of contaminants in soil and ground water. These in-situ 
processes include, biodegradation, dispersion, dilution, sorption, volatilization, and 
chemical or biological stabilization, transformation, or destruction of contaminants. 

Monitored natural attenuation is appropriate as a remedial approach only 
when it can be demonstrated capable of achieving a site’s remedial objectives within 
a time frame that is reasonable compared to that offered by other methods and 
where it meets the applicable remedy selection program for a particular OSWER 
program. EPA, therej'ore, expects that monitored natural attenution typically will 
be used in conjunction with active remediation measures (e.g., source control), or 
as a follow-up to active remediation measures that have already been implemented. 

The intent of this document is to present a technical protocol for data collection and analysis 
to evaluate monitored natural attenuation through biological processes for remediating ground 
water contaminated with mixtures of fuels and chlorinated aliphatic hydrocarbons. This document 
focuses on technical issues and is not intended to address policy considerations or specific regulatory 
or statutory requirements. In addition, this document does not provide comprehensive guidance on 
overall site characterization or long-term monitoring of MNA remedies. Users of this protocol 
should realize that different Federal and State remedial programs may have somewhat different 
remedial objectives. For example, the CERCLA and RCRA Corrective Action programs generally 
require that remedial actions: 1) prevent exposure to contaminated ground water, above acceptable 
risk levels; 2) minimize further migration of the plume; 3) minimize further migration of 
contaminants from source materials; and 4) restore the plume to cleanup levels appropriate for 
current or future beneficial uses, to the extent practicable. Achieving such objectives could often 
require that MNA be used in conjunction with other “active” remedial methods. For other cleanup 
programs, remedial objectives may be focused on preventing exposures above acceptable levels. 
Therefore, it is imperative that users of this document be aware of and understand the Federal and 


1 


State statutory and regulatory requirements, as well as policy considerations that apply to a specific 
site for which this protocol will be used to evaluate MNA as a remedial option. As a general 
practice (i.e., not just pertaining to this protocol), individuals responsible for evaluating remedial 
alternatives should interact with the overseeing regulatory agency to identify likely characterization 
and cleanup objectives for a particular site prior to investing significant resources. The policy 
framework within which MNA should be considered for Federal cleanup programs is described in 
the November 1997 EPA Directive titled, “Use of Monitored Natural Attenuation at Superfund, 
RCRA Corrective Action and Underground Storage Tank Sites” (Directive No. 9200.4-17). 

This protocol is designed to evaluate the fate in ground water of chlorinated aliphatic 
hydrocarbons and/or fuel hydrocarbons. Because documentation of natural attenuation requires 
detailed site characterization, the data collected under this protocol can be used to compare the 
relative effectiveness of other remedial options and natural attenuation. This protocol should be 
used to evaluate whether MNA by itself or in conjunction with other remedial technologies is 
sufficient to achieve site-specific remedial objectives. In evaluating the appropriateness of MNA, 
the user of this protocol should consider both existing exposure pathways, as well as exposure 
pathways arising from potential future uses of the ground water. 

This protocol is aimed at improving the characterization process for sites at which a remedy 
involving monitored natural attenuation is being considered. It contains methods and recommended 
strategies for completing the remedial investigation process. Emphasis is placed on developing a 
more complete understanding of the site through the conceptual site model process, early pathways 
analysis, and evaluation of remedial processes to include MNA. Understanding the contaminant 
flow field in the subsurface is essential for a technically justified evaluation of an MNA remedial 
option; therefore, use of this protocol is not appropriate for evaluating MNA at sites where the 
contaminant flow field cannot be determined with an acceptable degree of certainty (e.g., complex 
fractured bedrock, karst aquifers). 

In practice, natural attenuation also is referred to by several other names, such as intrinsic 
remediation, intrinsic bioremediation, natural restoration, or passive bioremediation. The goal of 
any site characterization effort is to understand the fate and transport of the contaminants of concern 
over time in order to assess any current or potential threat to human health or the environment. 
Natural attenuation processes, such as biodegradation, can often be dominant factors in the fate and 
transport of contaminants. Thus, consideration and quantification of natural attenuation is essential 
to a more thorough understanding of contaminant fate and transport. 

1.1 APPROPRIATE APPLICATION ON NATURAL ATTENUATION 

The intended audience for this document includes Project Managers and their contractors, 
scientists, consultants, regulatory personnel, and others charged with remediating ground water 
contaminated with chlorinated aliphatic hydrocarbons or mixtures of fuel hydrocarbons and 
chlorinated aliphatic hydrocarbons. This protocol is intended to be used within the established 
regulatory framework appropriate for selection of a remedy at a particular hazardous waste site 
(e.g., the nine-criteria analysis used to evaluate remedial alternatives in the CERCLA remedy 
selection process). It is not the intent of this document to replace existing U.S. EPA or state- 
specific guidance on conducting remedial investigations. 

The EPA does not consider monitored natural attenuation to be a default or presumptive 
remedy at any contaminated site (OSWER Directive 9200.4-17,1997), as its applicability is highly 
variable from site to site. In order for MNA to be selected as a remedy, site-specific determinations 


2 


will always have to be made to ensure that natural attenuation is sufficiently protective of human 
health and the environment. 

Natural attenuation in ground-water systems results from the integration of several subsurface 
attenuation mechanisms that are classified as either destructive or nondestructive. Biodegradation 
is the most important destructive attenuation mechanism, although abiotic destruction of some 
compounds does occur. Nondestructive attenuation mechanisms include sorption, dispersion, dilution 
from recharge, and volatilization. The natural attenuation of fuel hydrocarbons is described in the 
Technical Protocolfor Implementing Intrinsic Remediation with Long-Term Monitoring for Natural 
Attenuation of Fuel Contamination Dissolved in Groundwater, published by the Air Force Center 
for Environmental Excellence (AFCEE) (Wiedemeier et al ., 1995d). This document differs from 
the technical protocol for intrinsic remediation of fuel hydrocarbons because it focuses on the 
individual processes of chlorinated aliphatic hydrocarbon biodegradation which are fundamentally 
different from the processes involved in the biodegradation of fuel hydrocarbons. 

For example, biodegradation of fuel hydrocarbons, especially benzene, toluene, ethylbenzene, 
and xylenes (BTEX), is mainly limited by electron acceptor availability, and generally will proceed 
until all of the contaminants biochemically accessible to the microbes are destroyed. In the experience 
of the authors, there appears to be an adequate supply of electron acceptors in most, if not all, 
hydrogeologic environments. On the other hand, the more highly chlorinated solvents such as 
perchloroethene (PCE) and trichloroethene (TCE) typically are biodegraded under natural conditions 
via reductive dechlorination, a process that requires both electron acceptors (the chlorinated aliphatic 
hydrocarbons) and an adequate supply of electron donors. Electron donors include fuel hydrocarbons 
or other types of anthropogenic carbon (e.g., landfill leachate) or natural organic carbon. If the 
subsurface environment is depleted of electron donors before the chlorinated aliphatic hydrocarbons 
are removed, biological reductive dechlorination will cease, and natural attenuation may no longer 
be protective of human health and the environment. This is the most significant difference between 
the processes of fuel hydrocarbon and chlorinated aliphatic hydrocarbon biodegradation. 

For this reason, it is more difficult to predict the long-term behavior of chlorinated aliphatic 
hydrocarbon plumes than fuel hydrocarbon plumes. Thus, it is important to have a good 
understanding of the important natural attenuation mechanisms. Data collection should include all 
pertinent parameters to evaluate the efficacy of natural attenuation. In addition to having a better 
understanding of the processes of advection, dispersion, dilution from recharge, and sorption, it is 
necessary to better quantify biodegradation. This requires an understanding of the interactions 
between chlorinated aliphatic hydrocarbons, anthropogenic or natural carbon, and inorganic electron 
acceptors at the site. Detailed site characterization is required to adequately document and understand 
these processes. The long-term monitoring strategy should consider the possibility that the behavior 
of a plume may change over time and monitor for the continued availability of a carbon source to 
support reductive dechlorination. 

An understanding of the attenuation mechanisms is also important to characterizing exposure 
pathways. After ground water plumes come to steady state, sorption can no longer be an important 
attenuation mechanism. The most important mechanisms will be biotransformation, discharge 
through advective flow, and volatilization. As an example, Martin and Imbrigiotta (1994) calibrated 
a detailed transport and fate model to a release of pure TCE at Picatinny Arsenal, in New Jersey. 
The plume was at steady state or declining. Ten years after surface spills ceased, leaching of 
contaminants from subsurface DNAPLs and desorption from fine-grained layers were the only 
processes identified that continued to contribute TCE to ground water. Desorption of TCE occurred 


3 


at a rate of 15 to 85 mg/second. Anaerobic biotransformation consumed TCE at a rate of up to 30 
mg/second, advective flow and discharge of TCE to surface water accounted for up to 2 mg/ 
second, and volatilization of TCE accounted for 0.1 mg/second. In this case, recharge of 
uncontaminated water drove the plume below the water table, which minimized the opportunity for 
volatization to the unsaturated zone. As a result, discharge to surface water was the only important 
exposure pathway. Volatilization will be more important at sites that do not have significant 
recharge to the water table aquifer, or that have NAPLs at the water table that contain chlorinated 
organic compounds. 

Chlorinated solvents are released into the subsurface as either aqueous-phase or nonaqueous 
phase liquids. Typical solvent releases include nonaqueous phase relatively pure solvents that are 
more dense than water and aqueous rinseates. Additionally, a release may occur as a mixture of 
fuel hydrocarbons or sludges and chlorinated aliphatic hydrocarbons which, depending on the 
relative proportion of each compound group, may be more or less dense than water. If the NAPL 
is more dense than water, the material is referred to as a “dense nonaqueous-phase liquid,” or 
DNAPL. If the NAPL is less dense than water the material is referred to as a “light nonaqueous- 
phase liquid,” or LNAPL. Contaminant sources generally consist of chlorinated solvents present 
as mobile NAPL (NAPL occurring at sufficiently high saturations to drain under the influence of 
gravity into a well) and residual NAPL (NAPL occurring at immobile, residual saturations that are 
unable to drain into a well by gravity). In general, the greatest mass of contaminant is associated 
with these NAPL source areas, not with the aqueous phase. 

When released at the surface, NAPLs move downward under the force of gravity and tend to 
follow preferential pathways such as along the surface of sloping fine-grained layers or through 
fractures in soil or rock. Large NAPL releases can extend laterally much farther from the release 
point than would otherwise be expected, and large DNAPL releases can sink to greater depths than 
expected by following preferential flow paths. Thus, the relative volume of the release and potential 
migration pathways should be considered when developing the conceptual model for the distribution 
of NAPL in the subsurface. 

As water moves through NAPL areas (recharge in the vadose zone or ground water flow in an 
aquifer), the more soluble constituents partition into the water to generate a plume of dissolved 
contamination and the more volatile contaminants partition to the vapor phase. After surface 
releases have stopped, NAPLs remaining in the subsurface tend to “weather” over time as volatile 
and soluble components are depleted from NAPL surfaces. Even considering this “weathering” 
effect, subsurface NAPLS continue to be a source of contaminants to ground water for a very long 
time. For this reason, identification and delineation of subsurface zones containing residual or 
free-phase NAPL is an important aspect of the site conceptual model to be developed for evaluating 
MNA or other remediation methods. 

Removal, treatment or containment of NAPLs may be necessary for MNA to be a viable 
remedial option or to decrease the time needed for natural processes to attain site-specific remediation 
objectives. In cases where removal of mobile NAPL is feasible, it is desirable to remove this 
source material and decrease the time required to reach cleanup objectives. Where removal or 
treatment of NAPL is not practical, source containment may be practicable and necessary for MNA 
to be a viable remedial option. 

1.2 ADVANTAGES AND DISADVANTAGES 

In comparison to engineered remediation technologies, remedies relying on monitored natural 
attenuation have the following advantages and disadvantages, as identified in OSWER Directive 


4 


9200.4-17, dated November 1997. (Note that this an iterim, not a final, Directive which was released 
by EPA for use. Readers are cautioned to consult the final version of this Directive when it becomes 
available.) 

The advantages of monitored natural attenuation (MNA) remedies are: 

• As with any in situ process, generation of lesser volume ofremediation wastes reduced potential 
for cross-media transfer of contaminants commonly associated with ex situ treatment, and 
reduced risk of human exposure to contaminated media; 

• Less intrusion as few surface structures are required; 

• Potentialfor application to all or part of a given site, depending on site conditions and cleanup 
objectives; 

• Use in conjunction with, or as a follow-up to, other (active) remedial measures; and 

• Lower overall remediation costs than those associated with active remediation. 

The potential disadvantages of monitored natural attenuation (MNA) include: 

• Longer time frames may be required to achieve remediation objectives, compared to active 
remediation; 

• Site characterization may be more complex and costly; 

• Toxicity of transformation products may exceed that of the parent compound; 

• Long-term monitoring will generally be necessary; 

• Institutional controls may be necessary to ensure long-term protectiveness; 

• Potential exists for continued contamination migration, and/or cross-media transfer of 
contaminants; 

• Hydrologic and geochemical conditions amenable to natural attenuation are likely to change 
over time and could result in renewed mobility of previously stabilized contaminants, adversely 
impacting remedial effectiveness; and 

• More extensive education and outreach efforts may be required in order to gain public 
acceptance of monitored natural attenuation. 

At some sites the same geochemical conditions and processes that lead to biodegradation of 
chlorinated solvents and petroleum hydrocarbons can chemically transform naturally occurring 
manganese, arsenic and other metals in the aquifer matrix, producing forms of these metals that 
are more mobile and/or more toxic than the original materials. A comprehensive assessment of 
risk at a hazardous waste site should include sampling and analysis for these metals. 

This document describes (1) those processes that bring about natural attenuation, (2) the site 
characterization activities that may be performed to conduct a full-scale evaluation of natural 
attenuation, (3) mathematical modeling of natural attenuation using analytical or numerical solute 
fate and transport models, and (4) the post-modeling activities that should be completed to ensure 
successful evaluation and verification of remediation by natural attenuation. The objective is to 
quantify and provide defensible data to evaluate natural attenuation at sites where naturally occurring 
subsurface attenuation processes are capable of reducing dissolved chlorinated aliphatic hydrocarbon 
and/or fuel hydrocarbon concentrations to acceptable levels. A comment made by a member of the 
regulatory community summarizes what is required to successfully implement natural attenuation: 
A regulator looks for the data necessary to determine that a proposed 
treatment technology, if properly installed and operated, will reduce the contaminant 
concentrations in the soil and water to legally mandated limits. In this sense, the 
use of biological treatment systems calls for the same level of investigation, 


5 



demonstration of effectiveness, and monitoring as any conventional [remediation] 
system (National Research Council, 1993). 

When the rate of natural attenuation of site contaminants is sufficient to attain site-specific 
remediation objectives in a time period that is reasonable compared to other alternatives, MNA 
may be an appropriate remedy for the site. This document presents a technical course of action that 
allows converging lines of evidence to be used to scientifically document the occurrence of natural 
attenuation and quantify the rate at which it is occurring. Such a “weight-of-evidence” approach 
will greatly increase the likelihood of successfully implementing natural attenuation at sites where 
natural processes are restoring the environmental quality of ground water. 

1.3 LINES OF EVIDENCE 

The OSWER Directive 9200.4-17 (1997) identifies three lines of evidence that can be used to 
estimate natural attenuation of chlorinated aliphatic hydrocarbons, including: 

(1) Historical ground water and/or soil chemistry data that demonstrate a clear and meaningful 
trend of decreasing contaminant mass and/or concentration over time at appropriate 
monitoring or sampling points. (In the case of a ground water plume, decreasing 
concentrations should not be solely the result ofplume migration. In the case of inorganic 
contaminants, the primary attenuating mechanism should also be understood.) 

(2) Hydrogeologic and geochemical data that can be used to demonstrate indirectly the type(s) 
of natural attenuation processes active at the site, and the rate at which such processes 
will reduce contaminant concentrations to required levels. For example, characterization 
data may be used to quantify the rates of contaminant sorption, dilution, or volatilization, 
or to demonstrate and quantify the rates of biological degradation processes occurring at 
the site. 

(3) Data from field or microcosm studies (conducted in or with actual contaminated site 
media) which directly demonstrate the occurrence of a particular natural attenuation 
process at the site and its ability to degrade the contaminants of concern (typically used to 
demonstrate biological degradation processes only). 

The OSWER Directive provides the following guidance on interpreting the lines of evidence: 

Unless EPA or the implementing state agency determines that historical 
data (Number 1 above) are of sufficient quality and duration to support a decision 
to use monitored natural attenuation , EPA expects that data characterizing the 
nature and rates of natural attenuation processes at the site (Number 2 above) 
should be provided. Where the latter are also inadequate or inconclusive , data 
from microcosm studies (Number 3 above) may also be necessary. In general, 
more supporting information may be required to demonstrate the efficacy of 
monitored natural attenuation at those sites with contaminants which do not readily 
degrade through biological processes (e.g., most non-petroleum compounds, 
inorganics), at sites with contaminants that transform into more toxic and/or mobile 
forms than the parent contaminant, or at sites where monitoring has been performed 
for a relatively short period of time. The amount and type of information neededfor 
such a demonstration will depend upon a number of site-specific factors, such as the 
size and nature of the contamination problem, the proximity of receptors and the 
potential risk to those receptors, and other physical characteristics of the 
environmental setting (e.g., hydrogeology, ground cover, or climatic conditions). 


6 


The first line of evidence does not prove that contaminants are being destroyed. Reduction in 
contaminant concentration could be the result of advection, dispersion, dilution from recharge, 
sorption, and volatilization (i.e., the majority of apparent contaminant loss could be due to dilution). 
However, this line of evidence is critical for determining if any exposure pathways exist for current 
or potential future receptors. 

In order to evaluate remediation by natural attenuation at most sites, the investigator will 
have to determine whether contaminant mass is being destroyed. This is done using either, or 
both, of the second or third lines of evidence. The second line of evidence relies on chemical and 
physical data to show that contaminant mass is being destroyed, not just being diluted or sorbed to 
the aquifer matrix. For many contaminants, biodegradation is the most important process, but for 
certain contaminants nonbiological reactions are also important. The second line of evidence is 
divided into two components: 

• Using chemical analytical data in mass balance calculations to show that decreases in 
contaminant and electron acceptor/donor concentrations can be directly correlated to 
increases in metabolic end products/daughter compounds. This evidence can be used to 
show that electron acceptor/donor concentrations in ground water are sufficient to facilitate 
degradation of dissolved contaminants. Solute fate and transport models can be used to 
aid mass balance calculations and to collate and present information on degradation. 

• Using measured concentrations of contaminants and/or biologically recalcitrant tracers in 
conjunction with aquifer hydrogeologic parameters such as seepage velocity and dilution 
to show that a reduction in contaminant mass is occurring at the site and to calculate 
biodegradation rate constants. 

The biodegradation rate constants are used in conjunction with the other fate and transport 
parameters to predict contaminant concentrations and to assess risk at downgradient performance 
evaluation wells and within the area of the dissolved plume. 

Microcosm studies may be necessary to physically demonstrate that natural attenuation is 
occurring. Microcosm studies can also be used to show that indigenous biota are capable of degrading 
site contaminants at a particular rate. Microcosm studies for the purpose of developing rate 
constants should only be undertaken when they are the only means available to obtain biodegradation 
rate estimates. There are two important categories of sites where it is difficult or impossible to 
extract rate constants from concentrations of contaminants in monitoring wells in the field. In 
some sites, important segments of the flow path to receptors are not accessible to monitoring because 
of landscape features (such as lakes or rivers) or property boundaries that preclude access to a site 
for monitoring. In other sites that are influenced by tides, or the stage of major rivers, or ground 
water extraction wells, the ground water plume trajectory changes so rapidly that it must be described 
in a statistical manner. A “snapshot” round of sampling cannot be used to infer the plume velocity 
in calculations of the rate of attenuation. 

1.4 SITE CHARACTERIZATION 

The OSWER Directive 9200.4-17 (1997) describes EPA requirements for adequate site 
characterization. 

Decisions to employ monitored natural attenuation as a remedy or remedy 
component should be thoroughly and adequately supported with site-specific 
characterization data and analysis. In general, the level of site characterization 
necessary to support a comprehensive evaluation of natural attenuation is more 
detailed than that needed to support active remediation. Site characterizations for 


7 


natural attenuation generally warrant a quantitative understanding of source mass; 
ground water flow; contaminant phase distribution and partitioning between soil, 
ground water, and soil gas; rates of biological and non-biological transformation; 
and an understanding of how all of these factors are likely to vary with time. This 
information is generally necessary since contaminant behavior is governed by 
dynamic processes which must be well understood before natural attenuation can 
be appropriately applied at a site. Demonstrating the efficacy of this remediation 
approach likely will require analytical or numerical simulation of complex 
attenuation processes. Such analyses, which are critical to demonstrate natural 
attenuation’s ability to meet remedial action objectives, generally require a detailed 
conceptual site model as a foundation. 

A conceptual site model is a three-dimensional representation that conveys 
what is known or suspected about contamination sources, release mechanisms, and 
the transport and fate of those contaminants. The conceptual model provides the 
basis for assessing potential remedial technologies at the site. “Conceptual site 
model" is not synonymous with “computer model;" however, a computer model 
may be helpful for understanding and visualizing current site conditions or for 
predictive simulations of potential future conditions. Computer models, which 
simulate site processes mathematically, should in turn be based upon sound 
conceptual site models to provide meaningful information. Computer models typically 
require a lot of data, and the quality of the output from computer models is directly 
related to the quality of the input data. Because of the complexity of natural systems, 
models necessarily rely on simplifying assumptions that may or may not accurately 
represent the dynamics of the natural system. 

Site characterization should include collecting data to define (in three spatial 
dimensions over time) the nature and distribution of contamination sources as well 
as the extent of the ground water plume and its potential impacts on receptors. 
However, where monitored natural attenuation will be considered as a remedial 
approach, certain aspects of site characterization may require more detail or 
additional elements. For example, to assess the contributions of sorption, dilution, 
and dispersion to natural attenuation of contaminated ground water, a very detailed 
understanding of aquifer hydraulics, recharge and discharge areas and volumes, 
and chemical properties is required. Where biodegradation will be assessed, 
characterization also should include evaluation of the nutrients and electron donors 
and acceptors present in the ground water, the concentrations of co-metabolites 
and metabolic by-products, and perhaps specific analyses to identify the microbial 
populations present. The findings of these, and any other analyses pertinent to 
characterizing natural attenuation processes, should be incorporated into the 
conceptual model of contaminant fate and transport developed for the site. 

Development of an adequate database during the iterative site characterization process is an 
important step in the documentation of natural attenuation. Site characterization should provide 
data on the location, nature, phase distribution, and extent of contaminant sources. Site 
characterization also should provide information on the location, extent, and concentrations of 
dissolved contamination; ground water geochemical data; geologic information on the type and 
distribution of subsurface materials; and hydrogeologic parameters such as hydraulic conductivity, 


8 


hydraulic gradients, and potential contaminant migration pathways to human or ecological receptor 
exposure points. 

The data collected during site characterization can be used to simulate the fate and transport of 
contaminants in the subsurface. Such simulation allows prediction of the future extent and 
concentrations of the dissolved contaminant plume. Several types of models can be used to simulate 
dissolved contaminant transport and attenuation. 

The natural attenuation modeling effort has five primary objectives: 

• To evaluate whether MNA will be likely to attain site-specific remediation objectives in a 
time period that is reasonable compared to other alternatives; 

• To predict the future extent and concentration of a dissolved contaminant plume by 
simulating the combined effects of contaminant loading, advection, dispersion, sorption, 
and biodegradation; 

• To predict the most useful locations for ground-water monitoring; 

• To assess the potential for downgradient receptors to be exposed to contaminant 
concentrations that exceed regulatory or risk-based levels intended to be protective of 
human health and the environment; and 

• To provide technical support for remedial options using MNA during screening and detailed 
evaluation of remedial alternatives in a CERCLA Feasibility Study or RCRA Corrective 
Measures Study. 

Upon completion of the fate and transport modeling effort, model predictions can be used to 
evaluate whether MNA is a viable remedial alternative for a given site. If the transport and fate 
models predict that natural attenuation is sufficient to attain site-specific remediation objectives 
and will be protective of human health and the environment, natural attenuation may be an 
appropriate remedy for the site. Model assumptions and results should be verified by data obtained 
from site characterization. If model assumptions and results are not verified by site data, MNA is 
not likely to be a viable option and should not be proposed as the remedy. 

1.5 MONITORING 

The Monitoring Program OSWER Directive on Monitored Natural Attenuation (9200.4-17) 
describes EPA expectations for performance monitoring. 

Performance monitoring to evaluate remedy effectiveness and to ensure 
protection of human health and the environment is a critical element of all response 
actions. Performance monitoring is of even greater importance for monitored natural 
attenuation than for other types of remedies due to the longer remediation time 
frames, potential for ongoing contaminant migration, and other uncertainties 
associated with using monitored natural attenuation. This emphasis is underscored 
by EPA ’s reference to “monitored natural attenuation ”. 

The monitoring program developedfor each site should specify the location, 
frequency, and type of samples and measurements necessary to evaluate remedy 
performance as well as define the anticipated performance objectives of the remedy. 

In addition, all monitoring programs should be designed to accomplish the following: 

• Demonstrate that natural attenuation is occurring according to expectations; 

• Identify any potentially toxic transformation products resulting from 
biodegradation; 

• Determine if a plume is expanding (either downgradient, laterally or vertically); 

• Ensure no impact to downgradient receptors; 

• Detect new releases of contaminants to the environment that could impact the 


9 


effectiveness of the natural attenuation remedy; 

• Demonstrate the efficacy of institutional controls that were put in place to 
protect potential receptors; 

• Detect changes in environmental conditions (e.g., hydrogeologic, geochemical, 
microbiological, or other changes) that may reduce the efficacy of any of the 
natural attenuation processes; and 

• Verify attainment of cleanup objectives. 

Detection of changes will depend on the proper siting and constniction of 
monitoring wells/points. Although the siting of monitoring wells is a concern for 
any remediation technology, it is of even greater concern with monitored natural 
attenuation because of the lack of engineering controls to control contaminant 
migration. 

Performance monitoring should continue as long as contamination 
remains above required cleanup levels. Typically, monitoring is continued for a 
specified period (e.g., one to three years) after cleanup levels have been achieved to 
ensure that concentration levels are stable and remain below target levels. The 
institutional and financial mechanisms for maintaining the monitoring program 
should be clearly established in the remedy decision or other site documents, as 
appropriate. 

Natural attenuation is achieved when naturally occurring attenuation mechanisms, such as 
biodegradation, bring about a reduction in the total mass, toxicity, mobility, volume, or concentration 
of a contaminant dissolved in ground water. In some cases, natural attenuation processes will be 
capable of attaining site-specific remediation objectives in a time period that is reasonable compared 
to other alternatives. However, at this time, the authors are not aware of any sites where natural 
attenuation alone has succeeded in restoring ground water contaminated with chlorinated aliphatic 
hydrocarbons to drinking water quality over the entire plume. 

The material presented here was prepared through the joint effort between the Bioremediation 
Research Team at the Subsurface Protection and Remediation Division of U.S. EPA’s National 
Risk Management Research Laboratory (NRMRL) in Ada, Oklahoma, and the U.S. Air Force 
Center for Environmental Excellence, Technology Transfer Division, Brooks Air Force Base, 
Texas, and Parsons Engineering Science, Inc. (Parsons ES). It is designed to facilitate proper 
evaluation of remedial alternatives including natural attenuation at large chlorinated aliphatic 
hydrocarbon-contaminated sites. 

This information is the most current available at the time of this writing. The scientific 
knowledge and experience with natural attenuation of chlorinated solvents is growing rapidly and 
the authors expect that the process for evaluating natural attenuation of chlorinated solvents will 
continue to evolve. 

This document contains three sections, including this introduction. Section 2 presents the 
protocol to be used to obtain scientific data to evaluate the natural attenuation option. Section 3 
presents the references used in preparing this document. Appendix A describes the collection of 
site characterization data necessary to evaluate natural attenuation, and provides soil and ground- 
water sampling procedures and analytical protocols. Appendix B provides an in-depth discussion 
of the destructive and nondestructive mechanisms of natural attenuation. Appendix C covers data 
interpretation and pre-modeling calculations. 


10 


SECTION 2 

PROTOCOL FOR EVALUATING NATURAL ATTENUATION 


The primary objective of the natural attenuation investigation is to determine whether natural 
processes will be capable of attaining site-specific remediation objectives in a time period that is 
reasonable compared to other alternatives. Further, natural attenuation should be evaluated to 
determine if it can meet all appropriate Federal and State remediation objectives for a given site. 
This requires that projections of the potential extent of the contaminant plume in time and space be 
made. These projections should be based on historic variations in contaminant concentration, and 
the current extent and concentrations of contaminants in the plume in conjunction with measured 
rates of contaminant attenuation. Because of the inherent uncertainty associated with such predictions, 
it is the responsibility of the proponent of monitored natural attenuation to provide sufficient evidence 
to demonstrate that the mechanisms of natural attenuation will meet the remediation objectives 
appropriate for the site. This can be facilitated by using conservative parameters in solute fate and 
transport models and numerous sensitivity analyses'in order to better evaluate plausible contaminant 
migration scenarios. When possible, both historical data and modeling should be used to provide 
information that collectively and consistently confirms the natural reduction and removal of the 
dissolved contaminant plume. 

Figure 2.1 outlines the steps involved in a natural attenuation demonstration and shows the 
important regulatory decision points for implementing natural attenuation. For example, a Superfund 
Feasibility Study is a two-step process that involves initial screening of potential remedial alternatives 
followed by more detailed evaluation of alternatives that pass the screening step. A similar process 
is followed in a RCRA Corrective Measures Study and for sites regulated by State remediation 
programs. The key steps for evaluating natural attenuation are outlined in Figure 2.1 and include: 

1) Review available site data and develop a preliminary conceptual model. Determine if 
receptor pathways have already been completed. Respond as appropriate. 

2) If sufficient existing data of appropriate quality exist, apply the screening process de¬ 
scribed in Section 2.2 to assess the potential for natural attenuation. 

3) If preliminary site data suggest natural attenuation is potentially appropriate, perform 
additional site characterization to further evaluate natural attenuation. If all the recom¬ 
mended screening parameters listed in Section 2.2 have been collected and the screening 
processes suggest that natural attenuation is not appropriate based on the potential for 
natural attenuation, evaluate whether other processes can meet the cleanup objectives for 
the site (e.g., abiotic degradation or transformation, volatilization, or sorption) or select a 
remedial option other than MNA. 

4) Refine conceptual model based on site characterization data, complete pre-modeling 
calculations, and document indicators of natural attenuation. 

5) Simulate, if necessary, natural attenuation using analytical or numerical solute fate and 
transport models that allow incorporation of a biodegradation term. 

6) Identify potential receptors and exposure points and conduct an exposure pathways analy¬ 
sis. 

7) Evaluate the need for supplemental source control measures. Additional source control 
may allow MNA to be a viable remedial option or decrease the time needed for natural 
processes to attain remedial objectives. 


11 


Review Available Site Data 
If Site Data are Adequate 
Develop Preliminary Conceptual Model 


Screen the Site using the Procedure 
Presented in Figure 2.3 


Are no 

Screening Criteria 
Met? 


YES 

Does it* 

Appear That 

Natural Attenuation Alone' 
Will Meet Regulatory 
Criteria? 

YES 

Perform Site Characterization 
to Evaluate Natural Attenuation 


Gather any Additional Data 
Necessary to Complete 
the Screening of the Site 



„ YES 


Engineered Remediation Required, 


Implement Other Protocols 


Evaluate Use of 
Selected Additional 
Remedial Options 
Including Source 
Removal or Source 
Control Along with 
Natural Attenuation 


Excavation I 


Refine Concep 
Complete P 
Calcu 

tual Model and 

re-Modeling 

ations 

> 

r— ' 

Simulate Natural Attenuation 
Using Solute Fate and 
Transport Models 


. 

Verify Model Assumptions and 

Results with Site 
Characterization Data 

J 

, 

Use Results of Modeling and 
Site-Specific Information in an 
Exposure Pathways Analysis 


NAPL / 

Recovery / 




Hydraulic 

Containment 




Vacuum 

Dewatering 



Cometabolic 

Bioventing 




Air 

Sparging 






Reactive 

Barrier 


Enhanced 

Bioremediation 


Perform Site Characterization 
to Support Remedy Decision Making 


Assess Potential For 
Natural Attenuation 
With Remediation 
System Installed 


Refine Conceptual Model and 
Complete Pre-Modeling 
Calculations 


Simulate Natural Attenuation 
Combined with Remedial 
Option Selected Above 
Using Solute Transport Models 


Verify Model Assumptions and 
Results with Site 
Characterization Data 


Use Results of Modeling and 
Site-Specific Information in 
an Exposure Assessment 


Will Remediation 
Objectives Be Met 
Without Posing Unacceptable, 
RisksTo Potential 
Receptors ? 


YES 


NO 


Develop Draft Plan for 
Performance Evaluation 
Monitoring Wells and 
Long-Term Monitoring 


NO 


Does 

"Revised Remediation" 
Strategy Meet Remediation" 
Objectives Without Posing 
Unacceptable Risks 
To Potential 
Receptors ?, 


Determine Remedial 
Measures to be ' 
Combined with MNA 


YES 


Present Findings 
and Proposed 
Remedy in 
Feasibility Study 


Figure 2.1 Natural attenuation of chlorinated solvents flow chart. 


12 






















































































































8) Prepare a long-term monitoring and verification plan for the selected alternative. In some 
cases, this includes monitored natural attenuation alone, or in other cases in concert with 
supplemental remediation systems. 

9) Present findings of natural attenuation studies in an appropriate remedy selection docu¬ 
ment, such as a CERCLA Feasibility or RCRA Corrective Measures Study. The appropri¬ 
ate regulatory agencies should be consulted early in the remedy selection process to clarify 
the remedial objectives that are appropriate for the site and any other requirements that the 
remedy will be expected to meet. However, it should be noted that remedy requirements 
are not finalized until a decision is signed, such as a CERCLA Record of Decision or a 
RCRA Statement of Basis. 

The following sections describe each of these steps in more detail. 

2.1 REVIEW AVAILABLE SITE DATA AND DEVELOP PRELIMINARY 
CONCEPTUAL MODEL 

The first step in the natural attenuation investigation is to review available site-specific data. 
Once this is done, it is possible to use the initial site screening processes presented in Section 2.2 to 
determine if natural attenuation is a viable remedial option. A thorough review of these data also 
allows development of a preliminary conceptual model. The preliminary conceptual model will 
help identify any shortcomings in the data and will facilitate placement of additional data collection 
points in the most scientifically advantageous and cost-effective manner possible. 

The following site information should be obtained during the review of available data. 
Information that is not available for this initial review should be collected during subsequent site 
investigations when refining the site conceptual model, as described in Section 2.3. 

• Nature, extent, and magnitude of contamination: 

— Nature and history of the contaminant release: 

—Catastrophic or gradual release of NAPL ? 

—More than one source area possible or present ? 

—Divergent or coalescing plumes ? 

— Three-dimensional distribution of dissolved contaminants and mobile and residual 
NAPLs. Often high concentrations of chlorinated solvents in ground water are the result 
of landfill leachates, rinse waters, or ruptures of water conveyance pipes. For LNAPLs 
the distribution of mobile and residual NAPL will be used to define the dissolved plume 
source area. For DNAPLs the distribution of the dissolved plume concentrations, in addition 
to any DNAPL will be used to define the plume source area. 

— Ground water and soil chemical data. 

— Historical water quality data showing variations in contaminant concentrations both 
vertically and horizontally. 

— Chemical and physical characteristics of the contaminants. 

— Potential for biodegradation of the contaminants. 

— Potential for natural attenuation to increase toxity and/or mobility of natural occurring 
metals. 

• Geologic and hydrogeologic data in three dimensions (If these data are not available, they 
should be collected for the natural attenuation demonstration and for any other remedial 
investigation or feasibility study): 

— Lithology and stratigraphic relationships. 

— Grain-size distribution (gravels vs. sand vs. silt vs. clay). 


13 


Aquifer hydraulic conductivity (vertical and horizontal, effectiveness of aquitards, 
calculation of vertical gradients). 

Ground-water flow gradients and potentiometric or water table surface maps (over 
several seasons, if possible). 

- Preferential flow paths. 

Interactions between ground water and surface water and rates of infiltration/recharge. 

• Locations of potential receptor exposure points: 

Ground water production and supply wells, and areas that can be deemed a potential source 
of drinking water. 

Downgradient and crossgradient discharge points including any discharges to surface waters 
or other ecosystems. 

- Vapor discharge to basements and other confined spaces. 

In some cases, site-specific data are limited. If this is the case, all future site characterization 
activities should include collecting the data necessary to screen the site for the use of monitored 
natural attenuation as a potential site remedy. Much of the data required to evaluate natural attenuation 
can be used to design and evaluate other remedial measures. 

Available site characterization data should be used to develop a conceptual model for the site. 
This conceptual model is a three-dimensional representation of the source area as a NAPL or 
region of highly contaminated ground water, of the surrounding uncontaminated area, of ground 
water flow properties, and of the solute transport system based on available geological, biological, 
geochemical, hydrological, climatological, and analytical data for the site. Data on the contaminant 
levels and aquifer characteristics should be obtained from wells and boreholes which will provide 
a clear three-dimensional picture of the hydrologic and geochemical characteristics of the site. 
High concentrations of dissolved contaminants can be the result of leachates, rinse waters and 
rupture of water conveyance lines, and are not necessarily associated with NAPLs. 

This type of conceptual model differs from the conceptual site models commonly used by risk 
assessors that qualitatively consider the location of contaminant sources, release mechanisms, 
transport pathways, exposure points, and receptors. However, the conceptual model of the ground 
water system facilitates identification of these risk-assessment elements for the exposure pathways 
analysis. After development, the conceptual model can be used to help determine optimal placement 
of additional data collection points, as necessafy, to aid in the natural attenuation investigation and 
to develop the solute fate and transport model. Contracting and management controls must be 
flexible enough to allow for the potential for revisions to the conceptual model and thus the data 
collection effort. 

Successful conceptual model development involves: 

• Definition of the problem to be solved (generally the three dimensional nature, magnitude, 
and extent of existing and future contamination). 

• Identification of the core or cores of the plume in three dimensions. The core or cores contain 
the highest concentration of contaminants. 

• Integration and presentation of available data, including: 

- Local geologic and topographic maps, 

- Geologic data, 

- Hydraulic data, 

- Biological data, 

- Geochemical data, and 

- Contaminant concentration and distribution data. 


14 


• Determination of additional data requirements, including: 

- Vertical profiling locations, boring locations and monitoring well spacing in three dimensions, 

- A sampling and analysis plan (SAP), and 

- Any data requirements listed in Section 2.1 that have not been adequately addressed. 

Table 2.1 contains the recommended soil and ground water analytical methods for evaluating 

the potential for natural attenuation of chlorinated aliphatic hydrocarbons and/or fuel hydrocarbons. 
Any plan to collect additional ground water and soil quality data should include the analytes listed 
in this table. Table 2.2 lists the availability of these analyses and the recommended data quality 
requirements. Since required procedures for field sampling, analytical methods and data quality 
objectives vary somewhat among regulatory programs, the methods to be used at a particular site 
should be developed in collaboration with the appropriate regulatory agencies. There are many 
documents which may aid in developing data quality objectives (e.g.,U.S. EPA Order 5360.1 and 
U.S. EPA QA/G-4 Guidance for the Data Quality Objectives Process). 

2.2 INITIAL SITE SCREENING 

After reviewing available site data and developing a preliminary conceptual model, an 
assessment of the potential for natural attenuation must be made. As stated previously, existing 
data can be useful to determine if natural attenuation is capable of attaining site-specific remediation 
objectives in a time period that is reasonable compared to other alternatives. This is achieved by 
first determining whether the plume is currently stable or migrating and the future extent of the 
plume based on (1) contaminant properties, including volatility, sorptive properties, and 
biodegradability; (2) aquifer properties, including hydraulic gradient, hydraulic conductivity, porosity 
and concentrations of native organic material in the sediment (TOC), and (3) the location of the 
plume and contaminant source relative to potential receptor exposure points (i.e., the distance between 
the leading edge of the plume and the potential receptor exposure points). These parameters 
(estimated or actual) are used in this section to make a preliminary assessment of the effectiveness 
of natural attenuation in reducing contaminant concentrations. 

If, after completing the steps outlined in this section, it appears that natural attenuation will be 
a significant factor in contaminant removal and a viable remedial alternative, detailed site 
characterization activities that will allow evaluation of this remedial option should be performed. 
If exposure pathways have already been completed and contaminant concentrations exceed protective 
levels, or if such completion is likely, an engineered remedy is needed to prevent such exposures 
and should be implemented as an early action. For this case, MNA may still be appropriate to attain 
long-term remediation objectives for the site. Even so, the collection of data to evaluate natural 
attenuation can be integrated into a comprehensive remedial strategy and may help reduce the cost 
and duration of engineered remedial measures such as intensive source removal operations or pump- 
and-treat technologies. 

2.2.1 Overview of Chlorinated Aliphatic Hydrocarbon Biodegradation 

Because biodegradation is usually the most important destructive process acting to reduce 
contaminant concentrations in ground water, an accurate estimate of the potential for natural 
biodegradation is important to consider when determining whether ground water contamination 
presents a substantial threat to human health and the environment. This information also will be 
useful when selecting the remedial alternative that will be most cost effective at eliminating or 
abating these threats should natural attenuation alone not prove to be sufficient. 


15 


Table 2.1 Soil, Soil Gas, and Ground-water Analytical Methods to Evaluate the Potential for Natural Attenuation of Chlorinated Solvents or Fuel 
Hydrocarbons in Ground Water. Analyses other than those listed in this table may be required for regulatory compliance. 


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Table 2.1 (Continued) 


Field or 
Fixed-Base 

Laboratory 

Fixed-base 

Field 

Field 

Laboratory 

Sample Volume, 

Sample Container, 
Sample Preservation 

Collect up to 40 mL of 
water in a glass or 
plastic container; cool 
to 4°C. 

Collect up to 40 mL of 
water in a glass or 
plastic container; cool 

to 4°C. 

Read from oxygen 

meter. 

Measure using a flow¬ 

through cell or over¬ 
flow cell. 

Recommended 
Frequency of 
Analysis 

Each sampling 
round 

Each sampling 
round 

Each sampling 
round 

Each sampling 
round 

Data Use 

Substrate for anaerobic 
microbial respiration. 

Same as above. 

To determine if a well is 
adequately purged for 
sampling. 

Used to classify plume and 
to determine if reductive 
dechlorination is possible 
in the absence of 
anthropogenic carbon. 

Comments 

If this method is 
used for sulfate 
analysis, do not use 
the field method. 

Colorimetric, if this 
method is used for 
sulfate analysis, do 
not use the fixed- 
base laboratory 
method. 

Field only 

Laboratory 

Method/Reference 

IC method E300 

Hach method # 8051 

Field probe with direct 
reading meter. 

SW9060 








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Filter if turbidity gives a response from the photometer before addition of the reagents that is as large or larger than the specified minimum quantification limit. 























Over the past two decades, numerous laboratory and field studies have demonstrated that 
subsurface microorganisms can degrade a variety of chlorinated solvents (e.g., Bouwer et al ., 1981; 
Miller and Guengerich, 1982; Wilson and Wilson, 1985; Nelson et al., 1986; Bouwer and Wright, 
1988; Lee, 1988; Little et al., 1988; Mayer et al. , 1988; Arciero et al. , 1989; Cline and Delfino, 
1989; Freedman and Gossett, 1989; Folsom et al., 1990; Harker and Kim, 1990; Alvarez-Cohen 
and McCarty, 1991a, 1991b; DeStefano et al., 1991; Henry, 1991; McCarty et al., 1992; Hartmans 
and de Bont, 1992; McCarty and Semprini, 1994; Vogel, 1994). Whereas fuel hydrocarbons are 
biodegraded through use as a primary substrate (electron donor), chlorinated aliphatic hydrocarbons 
may undergo biodegradation under three different circumstances: intentional use as an electron 
acceptor; intentional use as an electron donor; or, through cometabolism where degradation of the 
chlorinated organic is fortuitous and there is no benefit to the microorganism. At a given site, one 
or all of these circumstances may pertain, although at many sites the use of chlorinated aliphatic 
hydrocarbons as electron acceptors appears to be most important under natural conditions. In this 
case, biodegradation of chlorinated aliphatic hydrocarbons will be an electron-donor-limited process. 
Conversely, biodegradation of fuel hydrocarbons is an electron-acceptor-limited process. 

In an uncontaminated aquifer, native organic carbon is used as an electron donor, and dissolved 
oxygen (DO) is used first as the prime electron acceptor. Where anthropogenic carbon (e.g., as fuel 
hydrocarbons) is present, it also will be used as an electron donor. After the DO is consumed, 
anaerobic microorganisms typically use additional electron acceptors (as available) in the following 
order of preference: nitrate, ferric iron oxyhydroxide, sulfate, and finally carbon dioxide. Evaluation 
of the distribution of these electron acceptors can provide evidence of where and how chlorinated 
aliphatic hydrocarbon biodegradation is occurring. In addition, because chlorinated aliphatic 
hydrocarbons may be used as electron acceptors or electron donors (in competition with other 
acceptors or donors), isopleth maps showing the distribution of these compounds and their daughter 
products can provide evidence of the mechanisms of biodegradation working at a site. As with 
BTEX, the driving force behind oxidation-reduction reactions resulting in chlorinated aliphatic 
hydrocarbon degradation is electron transfer. Although thermodynamically favorable, most of the 
reactions involved in chlorinated aliphatic hydrocarbon reduction and oxidation do not proceed 
abiotically. Microorganisms are capable of carrying out the reactions, but they will facilitate only 
those oxidation-reduction reactions that have a net yield of energy. 

2.2.1.1 Mechanisms of Chlorinated Aliphatic Hydrocarbon Biodegradation 

The following sections describe the biodegradation of those compounds that are most prevalent 
and whose behavior is best understood. 

2.2.1.1.1 Electron Acceptor Reactions (Reductive Dehalogenation) 

The most important process for the natural biodegradation of the more highly chlorinated 
solvents is reductive dechlorination. During this process, the chlorinated hydrocarbon is used as an 
electron acceptor, not as a source of carbon, and a chlorine atom is removed and replaced with a 
hydrogen atom. Figure 2.2 illustrates the transformation of chlorinated ethenes via reductive 
dechlorination. In general, reductive dechlorination occurs by sequential dechlorination from PCE 
to TCE to DCE to VC to ethene. Depending upon environmental conditions, this sequence may be 
interrupted, with other processes then acting upon the products. During reductive dechlorination, 
all three isomers of DCE can theoretically be produced. However, Bouwer (1994) reports that 
under the influence of biodegradation, cis- 1,2-DCE is a more common intermediate than trans- 1,2- 
DCE, and that 1,1-DCE is the least prevalent of the three DCE isomers when they are present as 
daughter products. Reductive dechlorination of chlorinated solvent compounds is associated with 


23 



Figure 2.2 Reductive dehalogenation of chlorinated ethenes. 


24 























the accumulation of daughter products and an increase in the concentration of chloride ions. Reductive 
dechlorination affects each of the chlorinated ethenes differently. Of these compounds, PCE is the 
most susceptible to reductive dechlorination because it is the most oxidized. Conversely, VC is the 
least susceptible to reductive dechlorination because it is the least oxidized of these compounds. 
As a result, the rate of reductive dechlorination decreases as the degree of chlorination decreases 
(Vogel and McCarty, 1985; Bouwer, 1994). Murray and Richardson (1993) have postulated that 
this rate decrease may explain the accumulation of VC in PCE and TCE plumes that are undergoing 
reductive dechlorination. Reductive dechlorination has been demonstrated under nitrate- and iron- 
reducing conditions, but the most rapid biodegradation rates, affecting the widest range of chlorinated 
aliphatic hydrocarbons, occur under sulfate-reducing and methanogenic conditions (Bouwer, 1994). 
Because chlorinated aliphatic hydrocarbon compounds are used as electron acceptors during reductive 
dechlorination, there must be an appropriate source of carbon for microbial growth in order for this 
process to occur (Bouwer, 1994). Potential carbon sources include natural organic matter, fuel 
hydrocarbons, or other anthropogenic organic compounds such as those found in landfill leachate. 

2.2.1.1.2 Electron Donor Reactions 

Murray and Richardson (1993) write that microorganisms are generally believed to be incapable 
of growth using PCE and TCE as a primary substrate (i.e., electron donor). However, under aerobic 
and some anaerobic conditions, the less oxidized chlorinated aliphatic hydrocarbons (e.g., VC) can 
be used as the primary substrate in biologically mediated oxidation-reduction reactions (McCarty 
and Semprini, 1994). In this type of reaction, the facilitating microorganism obtains energy and 
organic carbon from the degraded chlorinated aliphatic hydrocarbon. In contrast to reactions in 
which the chlorinated aliphatic hydrocarbon is used as an electron acceptor, only the least oxidized 
chlorinated aliphatic hydrocarbons can be used as electron donors in biologically mediated oxidation- 
reduction reactions. McCarty and Semprini (1994) describe investigations in which VC and 1,2- 
dichloroethane (DCA) were shown to serve as primary substrates under aerobic conditions. These 
authors also document that dichloromethane has the potential to function as a primary substrate 
under either aerobic or anaerobic environments. In addition, Bradley and Chapelle (1996) show 
evidence of mineralization of VC under iron-reducing conditions so long as there is sufficient 
bioavailable iron (III). Aerobic metabolism of VC may be characterized by a loss of VC mass and 
a decreasing molar ratio of VC to other chlorinated aliphatic hydrocarbon compounds. In addition, 
Klier et al. ( 1998) and Bradley and Chapelle (1997) show mineralization of DCE to carbon dioxide 
under aerobic, Fe(III) reducing, and methanogenic conditions, respectively. 

2.2.1.1.3 Cometabolism 

When a chlorinated aliphatic hydrocarbon is biodegraded via cometabolism, the degradation 
is catalyzed by an enzyme or cofactor that is fortuitously produced by the organisms for other 
purposes. The organism receives no known benefit from the degradation of the chlorinated aliphatic 
hydrocarbon. Rather, the cometabolic degradation of the chlorinated aliphatic hydrocarbon may in 
fact be harmful to the microorganism responsible for the production of the enzyme or cofactor 
(McCarty and Semprini, 1994). Cometabolism is best documented in aerobic environments, although 
it potentially could occur under anaerobic conditions. It has been reported that under aerobic 
conditions chlorinated ethenes, with the exception of PCE, are susceptible to cometabolic degradation 
(Murray and Richardson, 1993; Vogel, 1994; McCarty and Semprini, 1994). Vogel (1994) further 
elaborates that the rate of cometabolism increases as the degree of dechlorination decreases. During 
cometabolism, the chlorinated alkene is indirectly transformed by bacteria as they use BTEX or 


25 


another substrate to meet their energy requirements. Therefore, the chlorinated alkene does not 
enhance the degradation of BTEX or other carbon sources, nor will its cometabolism interfere with 
the use of electron acceptors involved in the oxidation of those carbon sources. 

2.2.1.2 Behavior of Chlorinated Solvent Plumes 

Chlorinated solvent plumes can exhibit three types of behavior depending on the amount of 
solvent, the amount of biologically available organic carbon in the aquifer, the distribution and 
concentration of natural electron acceptors, and the types of electron acceptors being used. Individual 
plumes may exhibit all three types of behavior in different portions of the plume. The different 
types of plume behavior are summarized below. 

2.2.1.2.1 Type 1 Behavior 

Type 1 behavior occurs where the primary substrate is anthropogenic carbon (e.g., BTEX or 
landfill leachate), and microbial degradation of this anthropogenic carbon drives reductive 
dechlorination. When evaluating natural attenuation of a plume exhibiting Type 1 behavior, the 
following questions must be answered: 

1) Is the electron donor supply adequate to allow microbial reduction of the chlorinated 
organic compounds? In other words, will the microorganisms “strangle” before they 
“starve” (i.e., will they run out of chlorinated aliphatic hydrocarbons used as electron 
acceptors before they run out of anthropogenic carbon used as the primary substrate)? 

2) What is the role of competing electron acceptors (e.g., dissolved oxygen, nitrate, iron (III) 
and sulfate)? 

3) Is VC oxidized, or is it reduced? 

Appendices B and C discuss what these questions mean and how they are answered. Type 1 
behavior results in the rapid and extensive degradation of the more highly-chlorinated solvents 
such as PCE, TCE, and DCE. 

2.2.1.2.2 Type 2 Behavior 

Type 2 behavior dominates in areas that are characterized by relatively high concentrations of 
biologically available native organic carbon. Microbial utilization of this natural carbon source 
drives reductive dechlorination (i.e., it is the primary substrate for microorganism growth). When 
evaluating natural attenuation of a Type 2 chlorinated solvent plume, the same questions as those 
posed in the description of Type 1 behavior must be answered. Type 2 behavior generally results in 
slower biodegradation of the highly chlorinated solvents than Type 1 behavior, but under the right 
conditions (e.g., areas with high natural organic carbon contents), this type of behavior also can 
result in rapid degradation of these compounds. 

2.2.1.23 Type 3 Behavior 

Type 3 behavior dominates in areas that are characterized by inadequate concentrations of 
native and/or anthropogenic carbon, and concentrations of dissolved oxygen that are greater than 
1.0 mg/L. Under these aerobic conditions, reductive dechlorination will not occur. The most 
significant natural attenuation mechanisms for PCE, TCE, and DCE will be advection, dispersion, 
and sorption. However, VC can be rapidly oxidized under these conditions. Type 3 behavior also 
occurs in ground water that does not contain microbes capable of biodegradation of chlorinated 
solvents. 


26 


2.2.1.2.4 Mixed Behavior 

As mentioned above, a single chlorinated solvent plume can exhibit all three types of behavior 
in different portions of the plume. This can be beneficial for natural biodegradation of chlorinated 
aliphatic hydrocarbon plumes. For example, Wiedemeier et al. (1996a) describe a plume at 
Plattsburgh AFB, New York, that exhibits Type 1 behavior in the source area and Type 3 behavior 
downgradient from the source. The most fortuitous scenario involves a plume in which PCE, TCE, 
and DCE are reductively dechlorinated with accumulation of VC near the source area (Type 1 or 
Type 2 behavior), then VC is oxidized (Type 3 behavior), either aerobically or via iron reduction 
further downgradient. Vinyl chloride is oxidized to carbon dioxide in this type of plume and does 
not accumulate. The following sequence of reactions occurs in a plume that exhibits this type of 
mixed behavior. 

PCE—>TCE—»DCE—>VC-»Carbon Dioxide 

In general, TCE, DCE, and VC may attenuate at approximately the same rate, and thus these 
reactions may be confused with simple dilution. Note that no ethene is produced during this reaction. 
Vinyl chloride is removed from the system much faster under these conditions than it is under VC- 
reducing conditions. 

A less desirable scenario, but one in which all contaminants may be entirely biodegraded, 
involves a plume in which all chlorinated aliphatic hydrocarbons are reductively dechlorinated via 
Type 1 or Type 2 behavior. Vinyl chloride is reduced to ethene, which may be further reduced to 
ethane or methane. The following sequence of reactions occurs in this type of plume. 

PCE-» TCE-> DCE->VC->Ethene—>Ethane 

This sequence has been investigated by Freedman and Gossett (1989). In this type of plume, 
VC degrades more slowly than TCE, and thus tends to accumulate. 

2.2.2 Bioattenuation Screening Process 

An accurate assessment of the potential for natural biodegradation of chlorinated compounds 
should be made before investing in a detailed study of natural attenuation. The screening process 
presented in this section is outlined in Figure 2.3. This approach should allow the investigator to 
determine if natural bioattenuation of PCE, TCE, DCE, TCA, and chlorobenzenes is likely to be a 
viable remedial alternative before additional time and money are expended. If the site is regulated 
under CERCLA, much of the data required to make the preliminary assessment of natural attenuation 
will be used to evaluate alternative engineered remedial solutions as required by the NCP. Table 2.3 
presents the analytical screening criteria. 

For most of the chlorinated solvents, the initial biotransformation in the environment is a 
reductive dechlorination. The initial screening process is designed to recognize geochemical 
environments where reductive dechlorination is plausible. It is recognized, however, that 
bioodegradation of certain halogenated compounds can also proceed via oxidative pathways. 
Examples include DCE, VC, the dichloroethanes, chloroethane, dichlorobenzenes, 
monochlorobenzene, methylene chloride, and ethylene dibromide. 

The following information is required for the screening process: 

• The chemical and geochemical data presented in Table 2.3 for background and target 
areas of the plume as depicted in Figure 2.4. Figure 2.4 shows the schematic locations of 
these data collection points. Note: If other contaminants are suspected, then data on the 
concentrations and distribution of these compounds also should be obtained. 

• Locations of source(s) and potential points of exposure. If subsurface NAPLs are 
sources, estimate extent of residual and free-phase NAPL. 

• An estimate of the transport velocity and direction of ground-water flow. 


27 



Figure 2.3 Initial screening process flow chart. 


28 

















































Table 2.3 Analytical Parameters and Weighting for Preliminary Screening for Anaerobic 
Biodegradation Processes 87 


Analysis 

Concentration in 
Most Contaminated 
Zone 

Interpretation 

Value 

Oxygen* 

<0.5 mg/L 

Tolerated, suppresses the reductive pathway at higher 
concentrations 

3 

Oxygen* 

>5 mg/L 

Not tolerated; however, VC may be oxidized aerobically 

-3 

Nitrate* 

<1 mg/L 

At higher concentrations may compete with reductive pathway 

2 

Iron II* 

>1 mg/L 

Reductive pathway possible; VC may be oxidized under Fe(lll)- 
reducing conditions 

3 

Sulfate* 

<20 mg/L 

At higher concentrations may compete with reductive pathway 

2 

Sulfide* 

>1 mg/L 

Reductive pathway possible 

3 

Methane* 

<0.5 mg/L 

VC oxidizes 

0 


>0.5 mg/L 

Ultimate reductive daughter product, VC Accumulates 

3 

Oxidation Reduction 

<50 millivolts (mV) 

Reductive pathway possible 

1 

Potential* (ORP) 
against Ag/AgCI 
electrode 

<-100mV 

Reductive pathway likely 

2 

pH* 

5 < pH < 9 

Optimal rangd for reductive pathway 

0 


5 > pH >9 

Outside optimal range for reductive pathway 

-2 

TOC 

> 20 mg/L 

Carbon and energy source; drives dechlorination; can be 
natural or anthropogenic 

2 

Temperature* 

> 20°C 

At T >20°C biochemical process is accelerated 

1 

Carbon Dioxide 

>2x background 

Ultimate oxidative daughter product 

1 

Alkalinity 

>2x background 

Results from interaction between CO 2 and aguifer minerals 

1 

Chloride* 

>2x background 

Daughter product of organic chlorine 

2 

Hydrogen 

>1 nM 

Reductive pathway possible, VC may accumulate 

3 

Hydrogen 

<1 nM 

VC oxidized 

0 

Volatile Fatty Acids 

> 0.1 mg/L 

Intermediates resulting from biodegradation of more complex 
compounds; carbon and energy source 

2 

BTEX* 

> 0.1 mg/L 

Carbon and energy source; drives dechlorination 

2 

Tetrachloroethene 


Material released 

0 

Trichloroethene* 


Material released 

Daughter product of PCE 

0 

DCE* 


Material released 

Daughter product of TCE 

If cis is > 80% of total DCE it is likely a daughter product 

1,1-DCE can be chemical reaction product of TCA 

0 

2^ 

VC* 


Material released 

Daughter product of DCE 

0 

1,1,1-Trichloroethane* 


Material released 

0 

DCA 


Daughter product of TCA under reducing conditions 

2 

Carbon Tetrachloride 


Material released 

0 

Chloroethane* 


Daughter product of DCA or VC under reducing conditions 

2 

Ethene/Ethane 

>0.01 mg/L 

Daughter product of VC/ethene 

2 


>0.1 mg/L 


3 

Chloroform 


Material released 

Daughter product of Carbon Tetrachloride 

0 

2 

Dichloromethane 


Material released 

Daughter product of Chloroform 

0 

2 


* Required analysis, a/ Points awarded only if it can be shown that the compound is a daughter product (i.e., not a constituent of the source 


NAPL). 


29 




































Once these data have been collected, the screening process can be undertaken. The following 
steps summarize the screening processes: 

1) Determine if biodegradation is occurring using geochemical data. If biodegradation is 
occurring, proceed to step 2. If it is not, assess the amount and types of data available. If 
data are insufficient to determine if biodegradation is occurring, collect supplemental data. 
If all the recommended screening parameters listed in section 2.2 have been collected and 
the screening processes suggest that natural attenuation is not appropriate, the screening 
processes are finished. Perform site characterization to evaluate other remediation alterna¬ 
tives. 

2) Determine ground-water flow and solute transport parameters from representative field 
data. Dispersivity and porosity may be estimated from literature but the hydraulic conduc¬ 
tivity and the ground-water gradient and flow direction must be determined from field 
data. The investigator should use the highest valid hydraulic conductivity measured at the 
site during the preliminary screening because solute plumes tend to follow the path of 
least resistance (i.e., highest hydraulic conductivity). This will give the “worst-case” 
estimate of the solute migration distance over a given period of time. Compare this 
“worst-case” estimate with the rate of plume migration determined from site characteriza¬ 
tion data. Determine what degree of plume migration is accepable or unacceptable with 
respect to site-specific remediation objectives. 

3) Locate source(s) and potential points of exposure. If subsurface NAPLs are sources, 
estimate extent of residual and free-phase NAPL. 

4) Estimate the biodegradation rate constant. Biodegradation rate constants can be estimated 
using a conservative tracer found commingled with the contaminant plume, as described 
in Appendix C and by Wiedemeier et al. (1996b). When dealing with a plume that con¬ 
tains chlorinated solvents, this procedure can be modified to use chloride as a tracer. Rate 
constants derived from microcosm studies can also be used when site specific field data 
are inadequate or inconclusive. If it is not possible to estimate the biodegradation rate 
using these procedures, then use a range of accepted literature values for biodegradation of 
the contaminants of concern. Appendix C presents a range of biodegradation rate con¬ 
stants for various compounds. Althougji literature values may be used to estimate 
biogradation rates in the bioattenuation screening process described in Section 2.2, litera¬ 
ture values should not be used in the later more detailed analysis of natural attenuation, 
described in Section 2.3. 

5) Compare the rate of transport to the rate of attenuation. 

Use analytical solutions or a screening model such as BIOSCREEN. 

6) Determine if screening criteria are met. 

Step 1: Determine if Biodegradation is Occurring 

The first step in the screening process is to sample or use existing data for the areas represented 
in Figure 2.4 and analyze them for the parameters listed in Table 2.3 (see also Section 2.3.2). These 
areas should include (1) the most contaminated portion of the aquifer (generally in the “source” 
area with NAPL or high concentrations of contaminants in ground water ; (2) downgradient from 
the source area but still in the dissolved contaminant plume; (3) downgradient from the dissolved 
contaminant plume; and (4) upgradient and lateral locations that are not impacted by the plume. 
Although this figure is a simplified two-dimensional representation of the features of a contaminant 
plume, real plumes are three-dimensional objects. The sampling should be conducted in accordance 
with Appendix A. 


30 



O Representative Sampling Location- 


Figure 2.4 Target areas for collecting screening data. Note that the number and location of monitoring 
wells will vary with the three dimensional complexity of the plume(s). 


The sample collected in the NAPL source area provides information as to the predominant 
terminal electron-accepting process at the source area. In conjunction with the sample collected in 
the NAPL source zone, samples collected in the dissolved plume downgradient from the NAPL 
source zone allow the investigator (1) to determine if the plume is degrading with distance along 
the flow path and (2) to determine the distribution of electron acceptors and donors and metabolic 
by-products along the flow path. The sample collected downgradient from the dissolved plume 
aids in plume delineation and allows the investigator to determine if metabolic byproducts are 
present in an area of ground water that has been remediated. The upgradient and lateral samples 
allow delineation of the plume and determination of background concentrations of the electron 
acceptors and donors. 

After these samples have been analyzed for the parameters listed in Table 2.3, the investigator 
should analyze the data to determine if biodegradation is occurring. The right-hand column of 
Table 2.3 contains scoring values that can be used as a test to assess the likelihood that biodegradation 
is occurring. This method relies on the fact that biodegradation will cause predictable changes in 
ground water chemistry. For example, if the dissolved oxygen concentration in the area of the 
plume with the highest contaminant concentration is less than 0.5 milligrams per liter (mg/L), 3 
points are awarded. Table 2.4 summarizes the range of possible scores and gives an interpretation 
for each score. If the score totals 15 or more points, it is likely that biodegradation is occurring, and 
the investigator should proceed to Step 2. 


31 







Table 2.4 Interpretation of Points Awarded During Screening Step 1 


Score 

Interpretation 

0 to 5 

Inadequate evidence for anaerobic biodegradation* of chlorinated organics 

6 to 14 

Limited evidence for anaerobic biodegradation* of chlorinated organics 

15 to 20 

Adequate evidence for anaerobic biodegradation* of chlorinated organics 

>20 

Strong evidence for anaerobic biodegradation* of chlorinated organics 

*reductive dechlorination 


The following two examples illustrate how Step 1 of the screening process is implemented. 
The site used in the first example is a former fire training area contaminated with chlorinated 
solvents mixed with fuel hydrocarbons. The presence of the fuel hydrocarbons appears to reduce 
the ORP of the ground water to the extent that reductive dechlorination is favorable. The second 
example contains data from a dry cleaning site contaminated only with chlorinated solvents. This 
site was contaminated with spent cleaning solvents that were dumped into a shallow dry well situated 
just above a well-oxygenated, unconfined aquifer with low organic carbon concentrations of dissolved 
organic carbon. 

Example 1: Strong Evidence for Anaerobic Biodegradation (Reductive Dechlorination) of 
Chlorinated Organics 


Analyte 

Concentration in Most Contaminated Zone 

Points Awarded 

Dissolved Oxygen 

0.1 mg/L 

3 

Nitrate 

0.3 mg/L 

~T~ 

Iron (11) 

10 mg/L 

~T~ 

Sulfate 

2 mg/L 

2 

Methane 

5 mg/L 

~T~ 

ORP 

-190 mV 

2 

Chloride 

3 times background 

2 

PCE (released) 

1,000 pg/L 

0 

TCE (none released) 

1,200 pg/L 

2 

cis-DCE (none released) 

500 pg/L 

2 

VC (none released) 


2 


Total Points Awarded 

23 Points 


In this example, the investigator can infer that biodegradation is likely occurring at the time of 
sampling and may proceed to Step 2. 


Example 2: Anaerobic Biodegradation (Reductive Dechlorination) Unlikely 

Analyte_Concentration in Most Contaminated Zone_Points Awarded 


Dissolved Oxygen 

3 mg/L 

-3 

Nitrate 

0.3 mg/L 

2 

Iron (II) 

Not Detected (ND) 

0 

Sulfate 

10 mg/L 

2 

Methane 

ND 

0 

ORP 

+ 100 mV 

0 

Chloride 

background 

0 

TCE (released) 

1,200 pg/L 

0 

cis-DCE (none released) 

ND 

0 

VC (none released) 

ND 

0 


Total Points Awarded 

1 Point 


32 

































In this example, the investigator can infer that biodegradation is probably not occurring or is 
occurring too slowly to contribute to natural attenuation at the time of the sampling. In this case, 
the investigator should evaluate whether other natural attenuation processes can meet the cleanup 
objectives for the site (e.g., abiotic degradation or transformation, volatilization or sorption) or 
select a remedial option other than MNA. 

Step 2: Determine Ground-water Flow and Solute Transport Parameters 

After it has been shown that biodegradation is occurring, it is important to quantify ground- 
water flow and solute transport parameters. This will make it possible to use a solute transport 
model to quantitatively estimate the concentration of the plume and its direction and rate of travel. 
To use an analytical model, it is necessary to know the hydraulic gradient and hydraulic conductivity 
for the site and to have estimates of porosity and dispersivity. It also is helpful to know the coefficient 
of retardation. Quantification of these parameters is discussed in detail in Appendix B. 

In order to make the modeling as accurate as possible, the investigator must have site-specific 
hydraulic gradient and hydraulic conductivity data. To determine the ground-water flow and solute 
transport direction, it is necessary to have at least three accurately surveyed wells in each 
hydrogeologic unit of interest at the site. The porosity and dispersivity are generally estimated 
using accepted literature values for the aquifer matrix materials containing the plume at the site. If 
the investigator has total organic carbon data for soil, it is possible to estimate the coefficient of 
retardation; otherwise, it is conservative to assume that the solute transport and ground-water 
velocities are the same. Techniques to collect these data are discussed in the appendices. 

Step 3: Locate Sources and Receptor Exposure Points 

To determine the length of flow for the predictive modeling to be conducted in Step 5, it is 
important to know the distance between the source of contamination, the leading edge along the 
core of the dissolved plume, and any potential downgradient or cross-gradient receptor exposure 
points. 

Step 4: Estimate the Biodegradation Rate 

Biodegradation is the most important process that degrades contaminants in the subsurface; 
therefore, the biodegradation rate is one of the most important model input parameters. 
Biodegradation of chlorinated aliphatic hydrocarbons can be represented as a first-order rate constant. 
Whenever possible, use site-specific biodegradation rates estimated from field data collected along 
the core of the plume. Calculation of site-specific biodegradation rates is discussed in Appendix C. 
If it is not possible to determine site-specific biodegradation rates, then literature values may be 
used in a sensitivity analysis (Table C.3.5). A useful approach is to start with average values, and 
then to vary the model input to predict “best-case” and “worst-case” scenarios. Estimated 
biodegradation rates can be used only after it has been shown that biodegradation is occurring (see 
Step 1). Although literature values may be used to estimate biodegradation rates in the bioattenuation 
screening process described in Section 2.2, additional site information should be collected to 
determine biodegradation rates for the site when refining the site conceptual model, as described in 
Section 2.3. Literature values should not be used during the more detailed analysis. 

Step 5: Compare the Rate of Transport to the Rate of Attenuation 

At this early stage in the natural attenuation demonstration, comparison of the rate of solute 
transport to the rate of attenuation is best accomplished using an analytical model. Several models 
are available. It is suggested that the decay option be first order for use in any of the models. 

The primary purpose of comparing the rate of transport to the rate of natural attenuation is to 
determine if natural attenuation processes will be capable of attaining site-specific remediation 
objectives in a time period that is reasonable compared to other alternatives (i.e., to quantitatively 


33 


estimate if site contaminants are attenuating at a rate fast enough to prevent further plume migration 
and restore the plume to appropriate cleanup levels). The analytical model BIOSCREEN can be 
used to determine whether natural attenuation processes will be capable of meeting site-specific 
remediation objectives at some distance downgradiant of a source. The numerical model BIOPLUME 
III can be used to estimate whether site contaminants are attenuating at a rate fast enough to restore 
the plume to appropriate cleanup levels It is important to perform a sensitivity analysis to help 
evaluate the confidence in the preliminary screening modeling effort. For the purposes of the 
screening effort , if modeling shows that the screening criteria are met, the investigator can proceed 
with the natural attenuation evaluation. 

Step 6: Determine if Screening Criteria are Met 

Before proceeding with the full-scale natural attenuation evaluation, the investigator should 
ensure that the answers to both of the following questions are “yes”: 

• Has the plume moved a shorter distance than would be expected based on the known (or 
estimated) time since the contaminant release and the contaminant velocity in ground 
water, as calculated from site-specific measurements of hydraulic conductivity and 
hydraulic gradient, and estimates of effective porosity and contaminant retardation? 

• Is it likely that site contaminants are attenuating at rates sufficient to meet remediation 
objectives for the site in a time period that is reasonable compared to other alternatives? 

If the answers to these questions are “yes,” then the investigator is encouraged to proceed with 
the full-scale natural attenuation demonstration. 

2.3 COLLECT ADDITIONAL SITE CHARACTERIZATION DATA TO EVALUATE 

NATURAL ATTENUATION AS REQUIRED 

It is the responsibility of the proponent to “make the case” for natural attenuation. Thus, a 
credible and thorough site assessment is necessary to document the potential for natural attenuation 
to meet cleanup objectives. As discussed in Section 2.1, review of existing site characterization 
data is particularly useful before initiating site characterization activities. Such review should 
allow identification of data gaps and guide the most effective placement of additional data collection 
points. 

There are two goals during the site characterization phase of a natural attenuation investigation. 
The first is to collect the data needed to determine if natural mechanisms of contaminant attenuation 
are occurring at rates sufficient to attain site-specific remediation objectives in a time period that is 
reasonable compared to other alternatives. The second is to provide sufficient site-specific data to 
allow prediction of the future extent and concentrations of a contaminant plume through solute fate 
and transport modeling. Thus, detailed site characterization is required to achieve these goals and 
to support this remedial option. Adequate site characterization in support of natural attenuation 
requires that the following site-specific parameters be determined: 

• Location, nature, and extent of contaminant source area(s) (i.e., areas containing mobile 
or residual NAPL or highly contaminated ground water). 

• Chemical properties (e.g., composition, solubility, volatility, etc.) of contaminant source 
materials. 

• The potential for a continuing source due to sewers, leaking tanks, or pipelines, or other 
site activity. 

• Extent and types of soil and ground-water contamination. 

• Aquifer geochemical parameters (Table 2.1). 


34 




• Regional hydrogeology, including: 

- Drinking water aquifers, and 

- Regional confining units. 

• Local and site-specific hydrogeology, including: 

- Local drinking water aquifers; 

- Location of industrial, agricultural, and domestic water wells; 

- Patterns of aquifer use (current and future); 

- Lithology; 

- Site stratigraphy, including identification of transmissive and nontransmissive units; 

- Potential pathways for NAPL migration (e.g., surface topography and dip of confining 

layers); 

- Grain-size distribution (sand vs. silt vs. clay); 

- Aquifer hydraulic conductivity; 

- Ground water hydraulic information; 

- Preferential flow paths; 

- Locations and types of surface water bodies; and 

- Areas of local ground water recharge and discharge. 

• Identification of current and future potential exposure pathways, receptors, and exposure 
points. 

Many chlorinated solvent plumes have enough three-dimensional expression to make it 
impossible for a single well to adequately describe the plume at a particular location on a map of 
the site. 

Figure 2.5 depicts a cross section of a hypothetical site with three-dimensional expression of 
the plume. A documented source exists in the capillary fringe just above the water table. Such 
sources are usually found by recovering, extracting, and analyzing core material. This material can 
be (1) a release of LNAPL containing chlorinated solvents; (2) a release of pure chlorinated solvents 
that has been entrapped by capillary interactions in the capillary fringe; or (3) material that has 
experienced high concentrations of solvents in solution in ground water, has sorbed the solvents, 
and now is slowly desorbing the chlorinated solvents. Recharge of precipitation through this source 
produces a plume that appears to dive into the aquifer as it moves away from the source. This effect 
can be caused by recharge of clean ground water above the plume as it moves downgradient of the 
source, by collection of the plume into more hydraulically conductive material at the bottom of 
aquifer, or by density differences between the plume and the unimpacted ground water. 

Below the first hydrologic unit there is a second unit that has fine-textured material at the top 
and coarse-textured material at the bottom of the unit. In the hypothetical site, the fine-textured 
material at the top of the second unit has inhibited downward migration of a DNAPL, causing it to 
spread laterally at the bottom of the first unit and form a second source of ground-water contamination 
in the first unit. Because DNAPL below the water table tends to exist as diffuse and widely extended 
ganglia rather than of pools filling all the pore space, it is statistically improbable that the material 
sampled by conventional core sampling will contain DNAPL. Because these sources are so difficult 
to sample, these sources are cryptic to conventional sampling techniques. 

At the hypothetical site, DNAPL has found a pathway past the fine-textured material and has 
formed a second cryptic source area at the bottom of the second hydrologic unit. Compare Figure 2.6. 
The second hydrological unit at the hypothetical site has a different hydraulic gradient than the first 
unit. As a result, the plume in the second unit is moving in a different direction than the plume in 
the first unit. Biological processes occurring in one hydrological unit may not occur in another; a 
plume may show Type 2 behavior in one unit and Type 3 behavior in another. 


35 


Figure 2.5 



A cross section through a hypothetical release, illustrating the three-dimensional character 
of the plumes that may develop from a release of chlorinated solvents. 



Figure 2.6 A stacked plan representation of the plumes that may develop from the hypothetical release 
depicted in Figure 2.5. Each plan representation depicts a separate plume that can 
originate from discrete source areas produced from the same release of chlorinated solvents. 


36 









































As a consequence, it is critical to sample and evaluate the three-dimensional character of the 
site with respect to (1) interaction of contaminant releases with the aquifer matrix material, (2) 
local hydological features that control development and migration of plumes, and (3) the geochemical 
interactions that favor bioattenuation of chlorinated solvents. 

The following sections describe the methodologies that should be implemented to allow 
successful site characterization in support of natural attenuation. 

2.3.1 Characterization of Soils and Aquifer Matrix Materials 

In order to adequately define the subsurface hydrogeologic system and to determine the three- 
dimensional distribution of mobile and residual NAPL that can act as a continuing source of ground- 
water contamination, credible and thorough soil characterization must be completed. As appropriate, 
soil gas data may be collected and analyzed to better characterize soil contamination in the vadose 
zone. Depending on the status of the site, this work may have been completed during previous 
remedial investigation work. The results of soils characterization will be used as input into a solute 
fate and transport model to help define a contaminant source term and to support the natural 
attenuation investigation. 

The purpose of sampling soil and aquifer matrix material is to determine the subsurface 
distribution of hydrostratigraphic units and the distribution of mobile and residual NAPL, as well 
as pore water that contains high concentrations of the contaminants in the dissolved phase. These 
objectives can be achieved through the use of conventional soil borings or direct-push methods 
(e.g., Geoprobe® or cone penetrometer testing), and through collection of soil gas samples. All 
samples should be collected, described, analyzed, and disposed of in accordance with local, State, 
and Federal guidance. Appendix A contains suggested procedures for sample collection. These 
procedures may require modification to comply with local, State, and Federal regulations or to 
accommodate site-specific conditions. 

The analytical methods to be used for soil, aquifer matrix material, and soil gas sample analyses 
is presented in Table 2.1. This table includes all of the parameters necessary to document natural 
attenuation, including the effects of sorption, volatilization, and biodegradation. Each analyte is 
discussed separately below. 

• Volatile Organic Compounds: Knowledge of the location, distribution, concentration, 
and total mass of contaminants sorbed to soils or present as mobile or immobile NAPL is 
required to calculate contaminant partitioning from NAPL into ground water. This 
information is useful to predict the long-term persistence of source areas. Knowledge of 
the diffusive flux of volatile organic compounds from NAPLs or ground water to the 
atmosphere or other identified receptor for vapors is required to estimate exposure of the 
human population or ecological receptors to contaminant vapors. If the flux of vapors 
can be compared to the discharge of the contaminants in ground water, the contribution of 
volatilization to natural attenuation of contamination in ground water can be documented. 

• Total Organic Carbon: Knowledge of the TOC content of the aquifer matrix is 
important for sorption and solute-retardation calculations. TOC samples should be 
collected from a background location in the stratigraphic horizon(s) where most 
contaminant transport is expected to occur. 

• Oxygen and Carbon Dioxide: Oxygen and carbon dioxide soil gas measurements can be 
used to identify areas in the unsaturated zone where biodegradation is occurring. This 
can be a useful and relatively inexpensive way to identify NAPL source areas, particularly 
when solvents are codisposed with fuels or greases (AFCEE, 1994). 


37 


• Fuel and Chlorinated Volatile Organic Compounds: Knowledge of the distribution of 
contaminants in soil gas can be used as a cost-effective way to estimate the extent of soil 
contamination. 

2.3.2 Ground-water Characterization 

To adequately determine the amount and three-dimensional distribution of dissolved 
contamination and to document the occurrence of natural attenuation, ground-water samples must 
be collected and analyzed. Biodegradation of organic compounds, whether natural or anthropogenic, 
brings about measurable changes in the chemistry of ground water in the affected area. By measuring 
these changes, it is possible to document and quantitatively evaluate the importance of natural 
attenuation at a site. 

Ground-water sampling is conducted to determine the concentrations and distribution of 
contaminants, daughter products, and ground-water geochemical parameters. Ground-water samples 
may be obtained from monitoring wells or with point-source sampling devices such as a Geoprobe®, 
Hydropunch®, or cone penetrometer. All ground-water samples should be collected, handled, and 
disposed of in accordance with local, State, and Federal guidelines. Appendix A contains suggested 
procedures for ground-water sample collection. These procedures may need to be modified to 
comply with local, State, and Federal regulations or to accommodate site-specific conditions. 

The analytical protocol for ground-water sample analysis is presented in Table 2.1. This 
analytical protocol includes all of the parameters necessary to delineate dissolved contamination 
and to document natural attenuation, including the effects of sorption and biodegradation. Data 
obtained from the analysis of ground water for these analytes is used to scientifically document 
natural attenuation and can be used as input into a solute fate and transport model. The following 
paragraphs describe each ground-water analytical parameter and the use of each analyte in the 
natural attenuation demonstration. 

2.3.2.1 Volatile and Semivolatile Organic Compounds 

These analytes are used to determine the type, concentration, and distribution of contaminants 
and daughter products in the aquifer. In many cases, chlorinated solvents are found commingled 
with fuels or other hydrocarbons. At a minimum, the volatile organic compound (VOC) analysis 
(Method SW8260A) should be used, with the addition of the trimethylbenzene isomers if fuel 
hydrocarbons are present or suspected. The'combined dissolved concentrations of BTEX and 
trimethylbenzenes should not be greater than about 30 mg/L for a JP-4 spill (Smith et al., 1981) or 
about 135 mg/L for a gasoline spill (Cline et al ., 1991; American Petroleum Institute, 1985). If 
these compounds are found in higher concentrations, sampling errors such as emulsification of 
LNAPL in the ground-water sample likely have occurred and should be investigated. 

Maximum concentrations of chlorinated solvents dissolved in ground water from neat solvents 
should not exceed their solubilities in water. Appendix B contains solubilities for common 
contaminants. If contaminants are found in concentrations greater than their solubilities, then 
sampling errors such as emulsification of NAPL in the ground-water sample have likely occurred 
and should be investigated. 

2.3.2.2 Dissolved Oxygen 

Dissolved oxygen is the most thermodynamically favored electron acceptor used by microbes 
for the biodegradation of organic carbon, whether natural or anthropogenic. Anaerobic bacteria 
generally cannot function at dissolved oxygen concentrations greater than about 0.5 mg/L and, 
hence, reductive dechlorination will not occur. This is why it is important to have a source of 
carbon in the aquifer that can be used by aerobic microorganisms as a primary substrate. During 


38 


aerobic respiration, dissolved oxygen concentrations decrease. After depletion of dissolved oxygen, 
anaerobic microbes will use nitrate as an electron acceptor, followed by iron (III), then sulfate, and 
finally carbon dioxide (methanogenesis). Each sequential reaction drives the ORP of the ground 
water downward into the range within which reductive dechlorination can occur. Reductive 
dechlorination is most effective in the ORP range corresponding to sulfate reduction and 
methanogenesis, but dechlorination of PCE and TCE also may occur in the ORP range associated 
with denitrification or iron (III) reduction. Dehalogenation of DCE and VC generally are restricted 
to sulfate reducing and methanogenic conditions. 

Dissolved oxygen measurements should be taken during well purging and immediately before 
and after sample acquisition using a direct-reading meter. Because most well purging techniques 
can allow aeration of collected ground-water samples, it is important to minimize the potential for 
aeration as described in Appendix A. 

23.2.3 Nitrate 

After dissolved oxygen has been depleted in the microbiological treatment zone, nitrate may 
be used as an electron acceptor for anaerobic biodegradation of organic carbon via denitrification. 
In order for reductive dechlorination to occur, nitrate concentrations in the contaminated portion of 
the aquifer must be less than 1.0 mg/L. 

23.2.4 Iron (II) 

In some cases, iron (III) is used as an electron acceptor during anaerobic biodegradation of 
organic carbon. During this process, iron (III) is reduced to iron (II), which may be soluble in water. 
Iron (II) concentrations can thus be used as an indicator of anaerobic degradation of fuel compounds, 
and vinyl chloride (see Section 2.2.1.1.2). Native organic matter may also support reduction of iron 
(II). Care must be taken when interpreting iron (II) concentrations because they may be biased low 
by reprecipitation as sulfides or carbonates. 

23.2.5 Sulfate 

After dissolved oxygen and nitrate have been depleted in the microbiological treatment zone, 
sulfate may be used as an electron acceptor for anaerobic biodegradation. This process is termed 
“sulfate reduction” and results in the production of sulfide. Concentrations of sulfate greater than 
20 mg/L may cause competitive exclusion of dechlorination. However, in many plumes with high 
concentrations of sulfate, reductive dechlorination still occurs. 

23.2.6 Methane 

During methanogenesis acetate is split to form carbon dioxide and methane, or carbon dioxide 
is used as an electron acceptor, and is reduced to methane. Methanogenesis generally occurs after 
oxygen, nitrate, and sulfate have been depleted in the treatment zone. The presence of methane in 
ground water is indicative of strongly reducing conditions. Because methane is not present in fuel, 
the presence of methane above background concentrations in ground water in contact with fuels is 
indicative of microbial degradation of hydrocarbons. Methane also is associated with spills of pure 
chlorinated solvents (Weaver et al ., 1996). It is not known if the methane comes from chlorinated 
solvent carbon or from native dissolved organic carbon. 

23.2.7 Alkalinity 

There is a positive correlation between zones of microbial activity and increased alkalinity. 
Increases in alkalinity result from the dissolution of rock driven by the production of carbon dioxide 
produced by the metabolism of microorganisms. Alkalinity is important in the maintenance of 
ground-water pH because it buffers the ground water system against acids generated during both 


39 


aerobic and anaerobic biodegradation. In the experience of the authors, biodegradation of organic 
compounds rarely, if ever, generates enough acid to impact the pH of the ground water. 

2.3.2.8 Oxidation-Reduction Potential 

The ORP of ground water is a measure of electron activity and is an indicator of the relative 
tendency of a solution to accept or transfer electrons. Oxidation-reduction reactions in ground 
water containing organic compounds (natural or anthropogenic) are usually biologically mediated, 
and, therefore, the ORP of a ground water system depends upon and influences rates of 
biodegradation. Knowledge of the ORP of ground water also is important because some biological 
processes operate only within a prescribed range of ORP conditions. 

ORP measurements can be used to provide real-time data on the location of the contaminant 
plume, especially in areas undergoing anaerobic biodegradation. Mapping the ORP of the ground 
water while in the field helps the field scientist to determine the approximate location of the 
contaminant plume. To map the ORP of the ground water while in the field, it is important to have 
at least one ORP measurement (preferably more) from a well located upgradient from the plume. 
ORP measurements should be taken during well purging and immediately before and after sample 
acquisition using a direct-reading meter. Because most well purging techniques can allow aeration 
of collected ground-water samples (which can affect ORP measurements), it is important to minimize 
potential aeration by using a flow-through cell as outlined in Appendix A. 

Most discussion of oxidation reduction potential expresses the potential as if it were measured 
against the standard hydrogen electrode. Most electrodes and meters to measure oxidation-reduction 
potential use the silver/silver chloride electrode (Ag/AgCl) as the reference electrode. This protocol 
uses the potential against the Ag/AgCl electrode as the screening potential, not Eh as would be 
measured against the standard hydrogen electrode. 

2.3.2.9 Dissolved Hydrogen 

In some ground waters, PCE and TCE appear to attenuate, although significant concentrations 
of DCE and VC do not accumulate. In this situation, it is difficult to distinguish between Type 3 
behavior where the daughter products are not produced, and Type 1 or Type 2 behavior where the 
daughter products are removed very rapidly. In cases like this, the concentration of hydrogen can 
be used to identify ground waters where reductive dechlorination is occurring. If hydrogen 
concentrations are very low, reductive dechlorination is not efficient and Type 3 behavior is indicated. 
If hydrogen concentrations are greater than approximately 1 nM, rates of reductive dechlorination 
should have environmental significance and Type 1 or Type 2 behavior would be expected. 

Concentrations of dissolved hydrogen have been used to evaluate redox processes, and thus 
the efficiency of reductive dechlorination, in ground-water systems (Lovley and Goodwin, 1988; 
Lovley et al ., 1994; Chapelle et al ., 1995). Dissolved hydrogen is continuously produced in anoxic 
ground-water systems by fermentative microorganisms that decompose natural and anthropogenic 
organic matter. This H 2 is then consumed by respiratory microorganisms that use nitrate, Fe(III), 
sulfate, or CO, as terminal electron acceptors. This continuous cycling of H, is called interspecies 
hydrogen transfer. Significantly, nitrate-, Fe(III)-, sulfate- and CO,-reducing (methanogenic) 
microorganisms exhibit different efficiencies in utilizing the H, that is being continually produced. 
Nitrate reducers are highly efficient H 2 utilizers and maintain very low steady-state H 2 concentrations. 
Fe(III) reducers are slightly less efficient and thus maintain somewhat higher H, concentrations. 
Sulfate reducers and methanogenic bacteria are progressively less efficient and maintain even higher 
H, concentrations. Because each terminal electron accepting process has a characteristic H, 
concentration associated with it, H 2 concentrations can be an indicator of predominant redox 


40 


processes. These characteristic ranges are given in Table 2.5. An analytical protocol for quantifying 
H 2 concentrations in ground water is given in Appendix A. 

Table 2.5 Range of Hydrogen Concentrations for a Given Terminal Electron-Accepting Process 


Terminal Electron 

Hydrogen (H 2 ) 

Accepting Process 

Concentration (nanomoles per liter) 

Denitrification 

<0.1 

Iron (III) Reduction 

0.2 to 0.8 

Sulfate Reduction 

1 to 4 

Reductive Dechlorination 

>1 

Methanogenesis 

5-20 


Oxidation-reduction potential (ORP) measurements are based on the concept of thermodynamic 
equilibrium and, within the constraints of that assumption, can be used to evaluate redox processes 
in ground water systems. The H 2 method is based on the ecological concept of interspecies hydrogen 
transfer by microorganisms and, within the constraints of that assumption, can also be used to 
evaluate redox processes. These methods, therefore, are fundamentally different. A direct comparison 
of these methods (Chapelle et al., 1996) has shown that ORP measurements were effective in 
delineating oxic from anoxic ground water, but that ORP measurements could not distinguish between 
nitrate-reducing, Fe(III)-reducing, sulfate-reducing, or methanogenic zones in an aquifer. In contrast, 
the H, method could readily distinguish between different anaerobic zones. For those sites where 
distinguishing between different anaerobic processes is important, H 2 measurements are an available 
technology for making such distinctions. At sites where concentrations of redox sensitive parameters 
such as dissolved oxygen, iron (II), sulfide, and methane are sufficient to identify operative redox 
processes, H 2 concentrations are not always required to identify redox zonation and predict 
contaminant behavior. 

In practice, it is preferable to interpret H 2 concentrations in the context of electron acceptor 
availability and the presence of the final products of microbial metabolism (Chapelle et al., 1995). 
For example, if sulfate concentrations in ground water are less than 0.5 mg/L, methane concentrations 
are greater than 0.5 mg/L, and H 2 concentrations are in the 5 to 20 nM range, it can be concluded 
with a high degree of certainty that methanogenesis is the predominant redox process in the aquifer. 
Similar logic can be applied to identifying denitrification (presence of nitrate, H 2 <0.1 nM), Fe(III) 
reduction (production of Fe(II), H 2 concentrations ranging from 0.2 to 0.8 nM), and sulfate reduction 
(presence of sulfate, production of sulfide, H 2 concentrations ranging from 1 to 4 nM). Reductive 
dechlorination in the field has been documented at hydrogen concentrations that support sulfate 
reduction or methanogenesis. If hydrogen concentrations are high enough to support sulfate reduction 
or methanogenesis, then reductive dechlorination is probably occurring, even if other geochemical 
indicators as scored in Table 2.3 do not indicate that reductive dechlorination is possible. 

2.3.2.10 pH, Temperature, and Conductivity 

Because the pH, temperature, and conductivity of a ground-water sample can change 
significantly within a short time following sample acquisition, these parameters must be measured 
in the field in unfiltered, unpreserved, “fresh” water collected by the same technique as the samples 
taken for dissolved oxygen and ORP analyses. The measurements should be made in a clean 


41 









container separate from those intended for laboratory analysis, and the measured values should be 
recorded in the ground-water sampling record. 

The pH of ground water has an effect on the presence and activity of microbial populations in 
ground water. This is especially true for methanogens. Microbes capable of degrading chlorinated 
aliphatic hydrocarbons and petroleum hydrocarbon compounds generally prefer pH values varying 
from 6 to 8 standard units. 

Ground-water temperature directly affects the solubility of dissolved gasses and other 
geochemical species. Ground-water temperature also affects the metabolic activity of bacteria. 

Conductivity is a measure of the ability of a solution to conduct electricity. The conductivity 
of ground water is directly related to the concentration of ions in solution; conductivity increases as 
ion concentration increases. 

2.3.2.11 Chloride 

Chlorine is the most abundant of the halogens. Although chlorine can occur in oxidation 
states ranging from Cl' to Cl +7 , the chloride form (Cl ) is the only form of major significance in 
natural waters (Hem, 1985). Chloride forms ion pairs or complex ions with some of the cations 
present in natural waters, but these complexes are not strong enough to be of significance in the 
chemistry of fresh water (Hem, 1985). Chloride ions generally do not enter into oxidation-reduction 
reactions, form no important solute complexes with other ions unless the chloride concentration is 
extremely high, do not form salts of low solubility, are not significantly adsorbed on mineral surfaces, 
and play few vital biochemical roles (Hem, 1985). Thus, physical processes control the migration 
of chloride ions in the subsurface. Kaufman and Orlob (1956) conducted tracer experiments in 
ground water, and found that chloride moved through most of the soils tested more conservatively 
(i.e., with less retardation and loss) than any of the other tracers tested. 

During biodegradation of chlorinated hydrocarbons dissolved in ground water, chloride is 
released into the ground water. This results in chloride concentrations in ground water in the 
contaminant plume that are elevated relative to background concentrations. Because of the neutral 
chemical behavior of chloride, it can be used as a conservative tracer to estimate biodegradation 
rates, as discussed in Appendix C. 

2.3.3 Aquifer Parameter Estimation 

Estimates of aquifer parameters are necessary to accurately evaluate contaminant fate and 
transport. 

2.3.3.1 Hydraulic Conductivity 

Hydraulic conductivity is a measure of an aquifer’s ability to transmit water, and is perhaps the 
most important aquifer parameter governing fluid flow in the subsurface. The velocity of ground 
water and dissolved contamination is directly related to the hydraulic conductivity of the saturated 
zone. In addition, subsurface variations in hydraulic conductivity directly influence contaminant 
fate and transport by providing preferential paths for contaminant migration. Estimates of hydraulic 
conductivity are used to determine residence times for contaminants and tracers, and to determine 
the seepage velocity of ground water. 

The most common methods used to quantify hydraulic conductivity are aquifer pumping tests 
and slug tests (Appendix A). Another method that may be used to determine hydraulic conductivity 
is the borehole dilution test. One drawback to these methods is that they average hydraulic properties 
over the screened interval. To help alleviate this potential problem, the screened interval of the test 
wells should be selected after consideration is given to subsurface stratigraphy. 


42 


Information about subsurface stratigraphy should come from geologic logs of continuous cores 
or from cone penetrometer tests. The rate of filling of a Hydropunch® can be used to obtain a 
rough estimate of the local hydraulic conductivity at the same time the water sample is collected. 
The results of pressure dissipation data from cone penetrometer tests can be used to supplement the 
results obtained from pumping tests and slug tests. It is important that the location of the aquifer 
tests be designed to collect information to delineate the range of hydraulic conductivity both vertically 
and horizontally at the site. 

2.3.3.1.1 Pumping Tests in Wells 

Pumping tests done in wells provide information on the average hydraulic conductivity of the 
screened interval, but not the most transmissive horizon included in the screened interval. In 
contaminated areas, the extracted ground water generally must be collected and treated, increasing 
the difficulty of such testing. In addition, a minimum 4-inch-diameter well is typically required to 
complete pumping tests in highly transmissive aquifers because the 2-inch submersible pumps 
available today are not capable of producing a flow rate high enough for meaningful pumping tests. 
In areas with fairly uniform aquifer materials, pumping tests can be completed in uncontaminated 
areas, and the results can be used to estimate hydraulic conductivity in the contaminated area. 
Pumping tests should be conducted in wells that are screened in the most transmissive zones in the 
aquifer. If pumping tests are conducted in wells with more than fifteen feet of screen, a down-hole 
flowmeter test can be used to determine the interval actually contributing to flow. 

2.3.3.1.2 Slug Tests in Wells 

Slug tests are a commonly used alternative to pumping tests. One commonly cited drawback to 
slug testing is that this method generally gives hydraulic conductivity information only for the area 
immediately surrounding the monitoring well. Slug tests do, however, have two distinct advantages 
over pumping tests: they can be conducted in 2-inch monitoring wells, and they produce no water. 
If slug tests are going to be relied upon to provide information on the three-dimensional distribution 
of hydraulic conductivity in an aquifer, multiple slug tests must be performed. It is not advisable to 
rely on data from one slug test in one monitoring well. Because of this, slug tests should be 
conducted at several zones across the site, including a test in at least two wells which are narrowly 
screened in the most transmissive zone. There should also be tests in the less transmissive zones to 
provide an estimate of the range of values present on the site. 

2.3.3.1.3 Downhole Flowmeter 

Borehole flowmeter tests are conducted to investigate the relative vertical distribution of 
horizontal hydraulic conductivity in the screened interval of a well or the uncased portion of a 
borehole. These tests can be done to identify any preferential flow pathways within the portion of 
an aquifer intersecting the test well screen or the open borehole. The work of Molz and Young 
(1993), Molz et al. (1994), Young and Pearson (1995), and Young (1995) describes the means by 
which these tests may be conducted and interpreted. 

In general, measurements of ambient ground-water flow rates are collected at several regularly 
spaced locations along the screened interval of a well. Next, the well is pumped at a steady rate, 
and the measurements are repeated. The test data may be analyzed using the methods described by 
Molz and Young (1993) and Molz et al. (1994) to define the relative distribution of horizontal 
hydraulic conductivity within the screened interval of the test well. Estimates of bulk hydraulic 
conductivity from previous aquifer tests can be used to estimate the absolute hydraulic conductivity 
distribution at the test well. 


43 


Using flowmeter test data, one may be able to more thoroughly quantify the three-dimensional 
hydraulic conductivity distribution at a site. This is important for defining contaminant migration 
pathways and understanding solute transport at sites with heterogeneous aquifers. Even at sites 
where the hydrogeology appears relatively homogeneous, such data may point out previously 
undetected zones or layers of higher hydraulic conductivity that control contaminant migration. In 
addition, ground-water velocities calculated from hydraulic head, porosity, and hydraulic conductivity 
data may be used to evaluate site data or for simple transport calculations. In these cases, it is also 
important to have the best estimate possible of hydraulic conductivity for those units in which the 
contaminants are migrating. 

2.3.3.2 Hydraulic Gradient 

The horizontal hydraulic gradient is the change in hydraulic head (feet of water) divided by the 
distance of ground-water flow between head measurement points. To accurately determine the 
hydraulic gradient, it is necessary to measure ground-water levels in all monitoring wells and 
piezometers at a site. Because hydraulic gradients can change over a short distance within an 
aquifer, it is essential to have as much site-specific ground-water elevation information as possible 
so that accurate hydraulic gradient calculations can be made. In addition, seasonal variations in 
ground-water flow direction can have a profound influence on contaminant transport. Sites in 
upland areas are less likely to be affected by seasonal variations in ground-water flow direction than 
low-elevation sites situated near surface water bodies such as rivers and lakes. 

To determine the effect of seasonal variations in ground-water flow direction on contaminant 
transport, quarterly ground-water level measurements should be taken over a period of at least one 
year. For many sites, these data may already exist. If hydraulic gradient data over a one-year period 
are not available, natural attenuation can still be implemented, pending an analysis of seasonal 
variation in ground-water flow direction. 

2.3.3.3 Processes Causing an Apparent Reduction in Total Contaminant Mass 

Several processes cause reductions in contaminant concentrations and apparent reductions in 
the total mass of contaminant in a system. Processes causing apparent reductions in contaminant 
mass include dilution, sorption, and hydrodynamic dispersion. In order to determine the mass of 
contaminant removed from the system, it is necessary to correct observed concentrations for the 
effects of these processes. This is done by incorporating independent assessments of these processes 
into the comprehensive solute transport model. The following sections give a brief overview of the 
processes that result in apparent contaminant reduction. Appendix B describes these processes in 
detail. 

Dilution results in a reduction in contaminant concentrations and an apparent reduction in the 
total mass of contaminant in a system due to the introduction of additional water to the system. The 
two most common causes of dilution (real or apparent) are infiltration and sampling from monitoring 
wells screened over large vertical intervals. Infiltration can cause an apparent reduction in 
contaminant mass by mixing unaffected waters with the contaminant plume, thereby causing dilution. 
Monitoring wells screened over large vertical distances may dilute ground-water samples by mixing 
water from clean aquifer zones with contaminated water during sampling. To avoid potential dilution 
during sampling, monitoring wells should be screened over relatively small vertical intervals (e.g. 
5 feet). Nested wells should be used to define the vertical extent of contamination in the saturated 
zone. Appendix C contains example calculations showing how to correct for the effects of dilution. 


44 


The retardation of organic solutes caused by sorption is an important consideration when 
simulating the effects of natural attenuation over time. Sorption of a contaminant to the aquifer 
matrix results in an apparent decrease in contaminant mass because dissolved contamination is 
removed from the aqueous phase. The processes of contaminant sorption and retardation are 
discussed in Appendix B. 

The dispersion of organic solutes in an aquifer is another important consideration when 
simulating natural attenuation. The dispersion of a contaminant into relatively pristine portions of 
the aquifer allows the solute plume to mix with uncontaminated ground water containing higher 
concentrations of electron acceptors. Dispersion occurs vertically as well as parallel and perpendicular 
to the direction of ground-water flow. 

To accurately determine the mass of contaminant transformed to innocuous by-products, it is 
important to correct measured contaminant concentrations for those processes that cause an apparent 
reduction in contaminant mass. This is accomplished by normalizing the measured concentration 
of each of the contaminants to the concentration of a tracer that is biologically recalcitrant. Because 
chloride is produced during the biodegradation of chlorinated solvents, this analyte can be used as 
a tracer. For chlorinated solvents undergoing reductive dechlorination, it is also possible to use the 
organic carbon in the original chlorinated solvent and daughter products as a tracer. Trimethylbenzene 
and tetramethylbenzene are two chemicals found in fuel hydrocarbon plumes that also may be 
useful as tracers. These compounds are difficult to biologically degrade under anaerobic conditions, 
and frequently persist in ground water longer than BTEX. Depending on the composition of the 
fuel that was released, other tracers may be used. 

2.3.4 Optional Confirmation of Biological Activity 

Extensive evidence can be found in the literature showing that biodegradation of chlorinated 
solvents and fuel hydrocarbons frequently occurs under natural conditions. Many references from 
the large body of literature in support of natural attenuation are listed in Section 3 and discussed in 
Appendix B. The most common technique used to show explicitly that microorganisms capable of 
degrading contaminants are present at a site is the microcosm study. 

If additional evidence (beyond contaminant and geochemical data and supporting calculations) 
supporting natural attenuation is required, a microcosm study using site-specific aquifer materials 
and contaminants can be undertaken. 

If properly designed, implemented, and interpreted, microcosm studies can provide very 
convincing documentation of the occurrence of biodegradation. Results of such studies are strongly 
influenced by the nature of the geological material submitted for study, the physical properties of 
the microcosm, the sampling strategy, and the duration of the study. Because microcosm studies 
are time-consuming and expensive, they should be undertaken only at sites where there is considerable 
uncertainty concerning the biodegradation of contaminants. 

Biodegradation rate constants determined by microcosm studies often are higher than rates 
achieved in the field. The collection of material for the microcosm study, the procedures used to set 
up and analyze the microcosm, and the interpretation of the results of the microcosm study are 
presented in Appendix C. 

2.4 REFINE CONCEPTUAL MODEL, COMPLETE PRE-MODELING CALCULA¬ 
TIONS, AND DOCUMENT INDICATORS OF NATURAL ATTENUATION 

Site investigation data should first be used to refine the conceptual model and quantify ground- 
water flow, sorption, dilution, and biodegradation. The results of these calculations are used to 
scientifically document the occurrence and rates of natural attenuation and to help simulate natural 


45 


attenuation over time. It is the responsibility of the proponent to “make the case” for natural 
attenuation. This being the case, all available data must be integrated in such a way that the evidence 
is sufficient to support the conclusion that natural attenuation is occurring. 

2.4.1 Conceptual Model Refinement 

Conceptual model refinement involves integrating newly gathered site characterization data to 
refine the preliminary conceptual model that was developed on the basis of previously collected 
site-specific data. During conceptual model refinement, all available site-specific data should be 
integrated to develop an accurate three-dimensional representation of the hydrogeologic and 
contaminant transport system. This refined conceptual model can then be used for contaminant 
fate and transport modeling. Conceptual model refinement consists of several steps, including 
preparation of geologic logs, hydrogeologic sections, potentiometric surface/water table maps, 
contaminant and daughter product contour (isopleth) maps, and electron acceptor and metabolic 
by-product contour (isopleth) maps. 

2.4.1.1 Geologic Logs 

Geologic logs of all subsurface materials encountered during the soil boring phase of the field 
work should be constructed. Descriptions of the aquifer matrix should include relative density, 
color, major and minor minerals, porosity, relative moisture content, plasticity of fines, cohesiveness, 
grain size, structure or stratification, relative permeability, and any other significant observations 
such as visible contaminants or contaminant odor. It is also important to correlate the results of 
VOC screening using soil sample headspace vapor analysis with depth intervals of geologic materials. 
The depth of lithologic contacts and/or significant textural changes should be recorded to the nearest 
0.1 foot. This resolution is necessary because preferential flow and contaminant transport paths 
may be limited to thin stratigraphic units. 

2.4.1.2 Cone Penetrometer Logs 

Cone Penetrometer Logs provide a valuable tool for the rapid collection of large amounts of 
stratigraphic information. When combined with the necessary corroborative physical soil samples 
from each stratigraphic unit occurring on the site, they can provide a three-dimensional model of 
subsurface stratigraphy. 

Cone penetrometer logs express stratigraphic information as the ratio of sleeve friction to tip 
pressure. Cone penetrometer logs also may cdntain fluid resistivity data and estimates of aquifer 
hydraulic conductivity. To provide meaningful data, the cone penetrometer must be capable of 
providing stratigraphic resolution on the order of 3 inches. To provide accurate stratigraphic 
information, cone penetrometer logs must be correlated with continuous subsurface cores. At a 
minimum, there must be one correlation for every hydrostratigraphic unit found at the site. Cone 
penetrometer logs, along with geologic boring logs, can be used to complete the hydrogeologic 
sections discussed in Section 2.4.1.3. 

2.4.1.3 Hydrogeologic Sections 

Hydrogeologic sections should be prepared from boring logs and/or CPT data. A minimum of 
two hydrogeologic sections are required; one parallel to the direction of ground-water flow and one 
perpendicular to the direction of ground water flow. More complex sites may require more 
hydrogeologic sections. Hydraulic head data including potentiometric surface and/or water table 
elevation data should be plotted on the hydrogeologic section. These sections are useful in identifying 
potential pathways of contaminant migration, including preferential pathways of NAPL migration 
(e.g., surface topography and dip of confining layers) and of aqueous contaminants (e.g., highly 


46 


transmissive layers). The potential distribution NAPL sources as well as preferential pathways for 
solute transport should be considered when simulating contaminant transport using fate and transport 
models. 

2.4.1.4 Potentiometric Surface or Water Table Map(s) 

A potentiometric surface or water table map is a two-dimensional graphic representation of 
equipotential lines shown in plan view. These maps should be prepared from water level 
measurements and surveyor’s data. Because ground water flows from areas of higher hydraulic 
head to areas of lower hydraulic head, such maps are used to estimate the probable direction of 
plume migration and to calculate hydraulic gradients. These maps should be prepared using water 
levels measured in wells screened in the same relative position within the same hydrogeologic unit. 
To determine vertical hydraulic gradients, separate potentiometric maps should be developed for 
different horizons in the aquifer to document vertical variations in ground-water flow. Flow nets 
should also be constructed to document vertical variations in ground-water flow. To document 
seasonal variations in ground-water flow, separate potentiometric surface or water table maps 
should be prepared for quarterly water level measurements taken over a period of at least one year. 
In areas with mobile LNAPL, a correction must be made for the water table deflection caused by 
accumlation of the LNAPL in the well. This correction and potentiometric surface map preparation 
are discussed in Appendix C. 

2.4.1.5 Contaminant and Daughter Product Contour Maps 

Contaminant and daughter product contour maps should be prepared for all contaminants 
present at the site for each discrete sampling event. Such maps allow interpretation of data on the 
distribution and the relative transport and degradation rates of contaminants in the subsurface. In 
addition, contaminant contour maps are necessary so that contaminant concentrations can be gridded 
and used for input into a numerical model. Detection of daughter products not present in the 
released NAPL (e.g., cw-l,2-DCE, VC, or ethene) provides evidence of reductive dechlorination. 
Preparation of contaminant isopleth maps is discussed in Appendix C. 

If mobile and residual NAPLs are present at the site, a contour map showing the thickness and 
vertical and horizontal distribution of each should be prepared. These maps will allow interpretation 
of the distribution and the relative transport rate of NAPLs in the subsurface. In addition, these 
maps will aid in partitioning calculations and solute fate and transport model development. It is 
important to note that, because of the differences between the magnitude of capillary suction in the 
aquifer matrix and the different surface tension properties of NAPL and water, NAPL thickness 
observations made at monitoring points may not provide an accurate estimate of the actual volume 
of mobile and residual NAPL in the aquifer. To accurately determine the distribution of NAPLs, it 
is necessary to take continuous soil cores or, if confident that chlorinated solvents present as NAPL 
are commingled with fuels, to use cone penetrometer testing coupled with laser-induced fluorescence. 
Appendix C discusses the relationship between actual and apparent NAPL thickness. 

2.4.1.6 Electron Acceptor, Metabolic By-product, and Alkalinity Contour Maps 

Contour maps should be prepared for electron acceptors consumed (dissolved oxygen, nitrate, 
and sulfate) and metabolic by-products produced [iron (II), chloride, and methane] during 
biodegradation. In addition, a contour map should be prepared for alkalinity and ORP. The electron 
acceptor, metabolic by-product, alkalinity, and ORP contour maps provide evidence of the occurrence 
of biodegradation at a site. If hydrogen data are available, they also should be contoured. 


47 


During aerobic biodegradation, dissolved oxygen concentrations will decrease to levels below 
background concentrations. Similarly, during anaerobic degradation, the concentrations of nitrate 
and sulfate will be seen to decrease to levels below background. The electron acceptor contour 
maps allow interpretation of data on the distribution of the electron acceptors and the relative transport 
and degradation rates of contaminants in the subsurface. Thus, electron acceptor contour maps 
provide visual evidence of biodegradation and a visual indication of the relationship between the 
contaminant plume and the various electron acceptors. 

Contour maps should be prepared for iron (II), chloride, and methane. During anaerobic 
degradation, the concentrations of these parameters will be seen to increase to levels above 
background. These maps allow interpretation of data on the distribution of metabolic by-products 
resulting from the microbial degradation of fuel hydrocarbons and the relative transport and 
degradation rates of contaminants in the subsurface. Thus, metabolic by-product contour maps 
provide visual evidence of biodegradation and a visual indication of the relationship between the 
contaminant plume and the various metabolic by-products. 

A contour map should be prepared for total alkalinity (as CaC0 3 ). Respiration of dissolved 
oxygen, nitrate, iron (III), and sulfate tends to increase the total alkalinity of ground water. Thus, 
the total alkalinity inside the contaminant plume generally increases to levels above background. 
This map will allow visual interpretation of alkalinity data by showing the relationship between the 
contaminant plume and elevated alkalinity. 

2.4.2 Pre-Modeling Calculations 

Several calculations must be made prior to implementation of the solute fate and transport 
model. These calculations include sorption and retardation calculations, NAPL/water partitioning 
calculations, ground-water flow velocity calculations, and biodegradation rate-constant calculations. 
Each of these calculations is discussed in the following sections. The specifics of each calculation 
are presented in the appendices referenced below. 

2.4.2.1 Analysis of Contaminant, Daughter Product, Electron Acceptor, Metabolic By-product, 
and Total Alkalinity Data 

The extent and distribution (vertical and horizontal) of contamination, daughter product, and 
electron acceptor and metabolic by-product concentrations are of paramount importance in 
documenting the occurrence of biodegradation <fnd in solute fate and transport model implementation. 

Comparison of contaminant, electron acceptor, electron donor, and metabolic by-product 
distributions can help identify significant trends in site biodegradation. Dissolved oxygen 
concentrations below background in an area with organic contamination are indicative of aerobic 
biodegradation of organic carbon. Similarly, nitrate and sulfate concentrations below background 
in an area with contamination are indicative of anaerobic biodegradation of organic carbon. Likewise, 
elevated concentrations of the metabolic by-products iron (II), chloride, and methane in areas with 
contamination are indicative of biodegradation of organic carbon. In addition, elevated concentrations 
of total alkalinity (as CaC0 3 ) in areas with contamination are indicative of biodegradation of organic 
compounds via aerobic respiration, denitrification, iron (III) reduction, and sulfate reduction. If 
these trends can be documented, it is possible to quantify the relative importance of each 
biodegradation mechanism, as described in Appendices B and C. The contour maps described in 
Section 2.4.1 can be used to provide graphical evidence of these relationships. 

Detection of daughter products not present in the released NAPL (e.g., cis-1,2-DCE, VC, or 
ethene) provides evidence of reductive dechlorination. The contour maps described in Section 2.4.1 
in conjunction with NAPL analyses can be used to show that reductive dechlorination is occurring. 


48 


2.4.2.2 Sorption and Retardation Calculations 

Contaminant sorption and retardation calculations should be made based on the TOC content 
of the aquifer matrix and the organic carbon partitioning coefficient (Koc) for each contaminant. 
The average TOC concentration from the most transmissive zone in the aquifer should be used for 
retardation calculations. A sensitivity analysis should also be performed during modeling using a 
range of TOC concentrations, including the lowest TOC concentration measured at the site. Sorption 
and retardation calculations should be completed for all contaminants and any tracers. Sorption 
and retardation calculations are described in Appendix C. 

2.4.2.3 NAPL/Water Partitioning Calculations 

If NAPL remains at the site, partitioning calculations should be made to account for the 
partitioning from this phase into ground water. Several models for NAPL/water partitioning have 
been proposed in recent years, including those by Hunt et al. (1988), Bruce et al. (1991), Cline et al. 
(1991), and Johnson and Pankow (1992). Because the models presented by Cline et al. ( 1991) and 
Bruce et al. (1991) represent equilibrium partitioning, they are the most conservative models. 
Equilibrium partitioning is conservative because it predicts the maximum dissolved concentration 
when NAPL in contact with water is allowed to re?xh equilibrium. The results of these equilibrium 
partitioning calculations can be used in a solute fate and transport model to simulate a continuing 
source of contamination. The theory behind fuel/water partitioning calculations is presented in 
Appendix B, and example calculations are presented in Appendix C. 

2.4.2.4 Ground-water Flow Velocity Calculations 

The average linear ground-water flow velocity of the most transmissive aquifer zone containing 
contamination should be calculated to check the accuracy of the solute fate and transport model and 
to allow calculation of first-order biodegradation rate constants. An example of a ground-water 
flow velocity calculation is given in Appendix C. 

2.4.2.5 Apparent Biodegradation Rate-Constant Calculations 

Biodegradation rate constants are necessary to accurately simulate the fate and transport of 
contaminants dissolved in ground water. In many cases, biodegradation of contaminants can be 
approximated using first-order kinetics. In order to calculate first-order biodegradation rate constants, 
the apparent degradation rate must be normalized for the effects of dilution, sorption, and 
volatilization. Two methods for determining first-order rate constants are described in Appendix C. 
One method involves the use of a biologically recalcitrant compound found in the dissolved 
contaminant plume that can be used as a conservative tracer. The other method, proposed by Buscheck 
and Alcantar (1995) is based on the one-dimensional steady-state analytical solution to the advection- 
dispersion equation presented by Bear (1979). It is appropriate for plumes where contaminant 
concentrations are in dynamic equilibrium between plume formation at the source and plume 
attenuation downgradient. Because of the complexity of estimating biodegradation rates with these 
methods, the results are more accurately referred to as “apparent” biodegradation rate constants. 
Apparent degradation rates reflect the difference between contaminant degradation and production 
which is important for some daughter products (e.g., TCE, DCE, and VC). 

2.5 SIMULATE NATURAL ATTENUATION USING SOLUTE FATE AND TRANS¬ 
PORT MODELS 

Simulating natural attenuation allows prediction of the migration and attenuation of the 
contaminant plume through time. Natural attenuation modeling is a tool that allows site-specific 
data to be used to predict the fate and transport of solutes under governing physical, chemical, and 


49 


biological processes. Hence, the results of the modeling effort are not in themselves sufficient 
proof that natural attenuation is occurring at a given site. The results of the modeling effort are only 
as good as the original data input into the model; therefore, an investment in thorough site 
characterization will improve the validity of the modeling results. In some cases, straightforward 
analytical models of solute transport are adequate to simulate natural attenuation. 

Several well-documented and widely accepted solute fate and transport models are available 
for simulating the fate and transport of contaminants under the influence of advection, dispersion, 
sorption, and biodegradation. 

2.6 CONDUCT A RECEPTOR EXPOSURE PATHWAYS ANALYSIS 

After the rates of natural attenuation have been documented, and predictions from appropriate 
fate and transport models indicate that MNA is a viable remedy, the proponent of natural attenuation 
should combine all available data and information to provide support for this remedial option. 
Supporting the natural attenuation option generally will involve performing a receptor exposure 
pathways analysis. This analysis includes identifying potential human and ecological receptors and 
points of exposure under current and future land and ground-water use scenarios. The results of 
solute fate and transport modeling are central to the exposure pathways analysis. If conservative 
model input parameters are used, the solute fate and transport model should give conservative 
estimates of contaminant plume migration. From this information, the potential for impacts on 
human health and the environment from contamination present at the site can be assessed. 

2.7 EVALUATE SUPPLEMENTAL SOURCE REMOVAL OPTIONS 

Additional source removal, treatment, or containment measures, beyond those previously 
implemented, may be necessary for MNA to be a viable remedial option or to decrease the time 
needed for natural processes to attain site-specific remedial objectives. Several technologies suitable 
for source reduction or removal are listed on Figure 2.1. Other technologies may be used as dictated 
by site conditions and regulatory requirements. If a solute fate and transport model has been prepared 
for a site, the impact of source removal can readily be evaluated by modifying the contaminant 
source term; this will allow for a reevaluation of the exposure pathways analysis. 

In some cases (particularly if the site is regulated under CERCLA), the removal, treatment, or 
containment of the source may be required to restore the aquifer as a source of drinking water, or to 
prevent discharge of contaminants to ecologically sensitive areas. If a solute fate and transport 
model has been prepared, it can also be used to forecast the benefits of source control by predicting 
the time required to restore the aquifer to drinking water quality, and the reduction in contaminant 
loadings to sensitive ecosystems. 

2.8 PREPARE LONG-TERM MONITORING PLAN 

This plan is used to monitor the plume over time and to verify that natural attenuation is 
occurring at rates sufficient to attain site-specific remediation objectives and within the time frame 
predicted at the time of remedy selection. In addition, the long-term monitoring plan should be 
designed to evaluate long-term behavior of the plume, verify that exposure to contaminants does 
not occur, verify that natural attenuation breakdown products do not pose additional risks, determine 
actual (rather than predicted) attenuation rates for refining predictions of remediation time frame, 
and to document when site-specific remediation objectives have been attained. 


50 


The long-term monitoring plan should be developed based on site characterization data, analysis 
of potential exposure pathways, and the results of solute fate and transport modeling. EPA is 
developing additional guidance on long-term monitoring of MNA remedies, which should be 
consulted when available. 

The long-term monitoring plan includes two types of monitoring wells. Long-term monitoring 
wells are intended to determine if the behavior of the plume is changing. Performance evaluation 
wells are intended to confirm that contaminant concentrations meet regulatory acceptance levels, 
and to trigger an action to manage potential expansion of the plume. Figure 2.7 depicts a schematic 
that describes the various categories of wells in a comprehensive monitoring plan. Figure 2.7 is 
intended to depict categories of wells, and does not depict monitoring well placement at a real site. 
Included in the schematic representation are: 1) wells in the source area; 2) wells in unimpacted 
ground water; 3) wells downgradient of the source area in a zone of natural attenuation; 4) wells 
located downgradient from the plume where contaminant concentrations are below regulatory 
acceptance levels but geochemical indicators are altered and soluble electron acceptors are depleted 
with respect to unimpacted ground water; and 5) performance evaluation wells. 

The final number and placement of long-term monitoring wells and performance evaluation 
wells will vary from site to site, based on the behavior of the plume as revealed during the site 
characterization and on the site-specific remediation objectives. In order to provide a valid monitoring 
system, all monitoring wells must be screened in the same hydrogeologic unit as the contaminant 
plume being monitored. This generally requires detailed stratigraphic correlation. To facilitate 
accurate stratigraphic correlation, detailed visual descriptions of all subsurface materials encountered 
during borehole drilling or cone penetrometer testing should be prepared prior to monitoring well 
installation. 



Figure 2.7 Hypothetical long-term monitoring strategy. Mote that number and location of monitoring 
wells will vary with the three-dimensional complexity of the plume(s) and site-specific 
remediation objectives. 


51 








Although the final number and placement of long-term monitoring wells and performance 
evaluation wells should be determined through regulatory negotiation, the locations of long-term 
monitoring wells should be based on the behavior of the plume as revealed during the site 
characterization and on regulatory considerations. The final number and location of performance 
evaluation wells will also depend on regulatory considerations. 

A ground-water sampling and analysis plan should be prepared in conjunction with a plan for 
placement of performance evaluation wells and long-term monitoring wells. For purposes of 
monitoring natural attenuation of chlorinated solvents, ground water from the long-term monitoring 
wells should be analyzed for the contaminants of concern, dissolved oxygen, nitrate, iron (II), sulfate, 
and methane. For performance evaluation wells, ground-water analyses should be limited to 
contaminants of concern. Any additional specific analytical requirements, such as sampling for 
contaminants that are metals, should be addressed in the sampling and analysis plan to ensure that 
all data required for regulatory decision making are collected. Water level and NAPL thickness 
measurements should be made during each sampling event. 

Except at sites with very low hydraulic conductivity and gradients, quarterly sampling of both 
long-term monitoring wells and performance evaluation wells is recommended during the first year 
to help determine whether the plume is stable or migrating, the direction of plume migration and to 
establish a baseline for behavior of the plume. After the first year, an appropriate sampling frequency 
should be established which considers seasonal variations in water table elevations, ground-water 
flow direction and flow velocity at the site. If the hydraulic conductivity or hydraulic gradient are 
low, the time required for ground water to move from upgradient monitoring wells to downgradient 
monitoring wells should also be considered in determining the appropriate monitoring frequency. 
Monitoring of long-term performance of an MNA remedy should continue as long as contamination 
remains above required cleanup levels. 

2.9 PRESENT FINDINGS 

Results of natural attenuation studies should be presented in the remedy selection document 
appropriate for the site, such as CERCLA Feasibility Study or RCRA Corrective Measures Study. 
This will provide scientific documentation that allows an objective evaluation of whether MNA is 
the most appropriate remedial option for a given site. 

All available site-specific data and information developed during the site characterization, 
conceptual model development, pre-modeling calculations, biodegradation rate calculation, ground- 
water modeling, model documentation, and long-term monitoring plan preparation phases of the 
natural attenuation investigation should be presented in a consistent and complementary manner in 
the feasibility study or similar document. Of particular interest to the site decision makers will be 
evidence that natural attenuation is occurring at rates sufficient to attain site-specific remediation 
objectives in a time period that is reasonable compared to other alternatives, and that human health 
and the environment will be protected over time. Since a weight-of-evidence argument will be 
presented to support an MNA remedy, all model assuptions should be conservative and all available 
evidence in support of MNA should be presented. 


52 


SECTION 3 
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53 


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' * 

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78 


APPENDIX A 


FIELD INVESTIGATION METHODOLOGIES 

> 


TABLE OF CONTENTS - APPENDIX A 


A-1 INTRODUCTION.A1 -4 

A-2 SUBSURFACE INVESTIGATION METHODOLOGIES. A2-5 

A.2.1 TRADITIONAL DRILLING TECHNIQUES.A2-5 

A.2.2 CONE PENETROMETER TESTING.A2-6 

A.2.3 HYDRAULIC PERCUSSION SYSTEMS.A2-7 

A-3 SOIL CHARACTERIZATION METHODOLOGIES.A3-8 

A.3.1 SAMPLE ACQUISITION.A3-8 

A.3.2 PHYSICAL DESCRIPTION.A3-8 

A.3.3 FIXED-BASE LABORATORY ANALYSES.A3-9 

A-4 GROUND-WATER CHARACTERIZATION METHODOLOGIES.A4-10 

A.4.1 GROUND-WATER MONITORING LOCATIONS, DEPTHS, AND SCREENED 

INTERVALS .A4-10 

A.4.2 TYPES OF GROUND-WATER SAMPLING LOCATIONS.A4-10 

A.4.2.1 Monitoring Wells.A4-11 

A.4.2.2 Monitoring Points.A4-11 

A.4.2.3 Grab Sampling.A4-12 

A.4.3 MEASUREMENT OF STATIC FLUID LEVELS.A4-12 

A.4.3.1 Water Level and Total Depth Measurements.A4-12 

A.4.3.2 Mobile LNAPL Thickness Measurements.A4-13 

A.4.3.4 Mobile DNAPL Thickness Measurements.A4-13 

A.4.3 GROUND-WATER EXTRACTION.A4-13 

A.4.3.1 Methods.A4-13 

A.4.3.2 Development.A4-14 

A.4.3.3 Purging.A4-15 

A.4.3.4 Sampling.A4-16 

A.4.4 GROUND-WATER ANALYTICAL PROCEDURES.A4-17 

A.4.4.1 Standard Well Head Analyses ' .A4-18 

A.4.4.2 Dissolved Hydrogen Analysis.A4-18 

A.4.4.2.1 Sampling Method.A4-18 

A.4.4.2.2 Analytical Method.A4-19 

A.4.4.3 Field Analytical Laboratory Analyses.A4-20 

A.4.4.4 Fixed-Base Laboratory Analyses.A4-22 

A-5 SURFACE WATER AND SEDIMENT CHARACTERIZATION 

METHODOLOGIES.A5-23 

A.5.1 Surface Water Sample Collection.A5-23 

A.5.2 Sediment Sample Collection.A5-23 

A-6 SAMPLE HANDLING.A6-24 

A.6.1 SAMPLE PRESERVATION, CONTAINERS, AND LABELS.A6-24 

A.6.2 SAMPLE SHIPMENT.A6-24 

A.6.3 CHAIN-OF-CUSTODY CONTROL.A6-24 

A.6.4 SAMPLING RECORDS. A6-25 


Al-2 









































A-7 AQUIFER CHARACTERIZATION METHODOLOGIES.A7-26 

A.7.1 HYDRAULIC CONDUCTIVITY.A7-26 

A.7.1.1 Pump Tests.A7-26 

A.7.1.1.1 Pumping Test Design.A7-26 

A.7.1.1.2 Preparation for Testing.A7-27 

A.7.1.1.3 Conducting the Pumping Test.A7-28 

A.7.1.2 Slug Tests.A7-29 

A.7.1.3 Downhole Flow Meter Measurements.A7-30 

A.7.2 HYDRAULIC GRADIENT.A7-31 

A.7.3 DIRECT MEASUREMENT OF GROUND-WATER VELOCITY. A7-31 

FIGURES 

No. Title Page 

A.4.1 Overflow Cell to Prevent Alteration of Geochemical 

Properties of Ground Water by Exposurp to the Atmosphere.A4-17 

A.4.2 Flowthrough Cell to Prevent Alteration of Geochemical 

Properties of Ground Water by Exposure to the Atmosphere.A4-17 

A.4.3 Schematic Showing the “Bubble Strip” Method for Measuring 

Dissolved Hydrogen Concentrations in Ground water.A4-19 


Al-3 















SECTION A-l 
INTRODUCTION 

Detailed site characterization is an important aspect of the remediation by monitored natural 
attenuation. Typically, it is necessary to collect additional site-specific data in order to successfully 
complete the demonstration. This appendix presents an overview of field techniques that can be 
used to collect the data used to evaluate monitored natural attenuation. These techniques are most 
appropriate for aquifers in unconsolidated sediments. They are less appropriate for fractured rock, 
and karst hydrogeologic settings. Selection of locations for field investigation activities and analyti¬ 
cal protocols used for soil and water samples are discussed in Section 2 of the protocol document. 

During all field investigation activities, special care should be taken to prevent contamination of 
the sampled matrices. The primary way that sample contamination can occur is through contact with 
improperly cleaned equipment. To prevent such contamination, proper equipment decontamination 
procedures must be developed and followed. Procedures will vary according to site contaminants, 
equipment type, field activity, sample matrix, rinseate handling requirements, and regulatory require¬ 
ments. All equipment requires decontamination prior to initiation of site activities and between 
sampling locations. New, disposable equipment does not require decontamination if factory-sealed 
and found acceptable according to the appropriate data quality objectives and the site specific Qual¬ 
ity Assurance Plan. In addition to the use of properly cleaned equipment, new, clean, disposable 
gloves (of a material appropriate to the activity and contaminant type/concentration) should be worn 
at each new sampling location. 

Basic health and safety precautions are required for every piece of equipment and every meth¬ 
odology discussed in this section. It is the responsibility of the investigator to be aware of and to 
communicate all health and safety issues to the field team; therefore, a site specific health and safety 
plan must be developed prior to initiating investigation activities. At a minimum this plan must 
contain: 

• A safety and health risk analysis for chemical, physical, and biological hazards associated 
with the site conditions, anticipated contaminants, equipment, field activities, and climate; 

• An emergency response plan with applicable emergency response numbers; and 

• Precautionary measures to be implemented to insure the safety of site workers. 

This appendix consists of seven sections, including this introduction. Section A-2 discusses 
subsurface investigation methodologies. Section A-3 discusses soil characterization methodologies. 
Section A-4 discusses groundwater characterization methodologies. Section A-5 discusses surface 
water and sediment characterization methodologies. Section A-6 discusses sample handling proce¬ 
dures. Section A-7 discusses aquifer characterization methodologies. 


A1-4 


SECTION A-2 

SUBSURFACE INVESTIGATION METHODOLOGIES 

The ideal technologies for an investigation of monitored natural attenuation are those which can 
rapidly provide a large amount of information in a very short period of time while producing low 
quantities of waste. The following subsections briefly introduce several alternatives that are avail¬ 
able for performing subsurface investigations to evaluate remediation by monitored natural attenua¬ 
tion. Although some of these alternatives more closely achieve the objectives of remediation by 
monitored natural attenuation investigation than others, considerations such as site geology, site 
hydrogeology, future well use, or regulatory concerns may dictate the selection of the subsurface 
investigation method for any given site. It is crucial to the evaluation of monitored natural attenua¬ 
tion to consider all of these issues prior to selecting a technology appropriate for their site. If during 
the investigation it becomes necessary to change methodologies, the same concerns must be re¬ 
addressed. 

Prior to initiating any intrusive subsurface activities, proposed drilling locations must be 
cleared. It is particularly useful if all utility lines in the investigation area are marked should changes 
to the investigation become necessary. In addition, in order to expedite the investigation, all neces¬ 
sary digging, coring, drilling, and ground-water monitoring point installation permits should be 
obtained prior to mobilizing to the field. Care should be taken not to cross-contaminate deeper 
aquifers by drilling through an aquitard underlying a DNAPL. 

At the conclusion of subsurface investigations, each sampling location that is not used to install 
a ground-water monitoring point or well should be restored as closely to its original condition as 
possible. Where possible, holes should be sealed with bentonite chips, pellets, or grout to eliminate 
any creation or enhancement of contaminant migration pathways to the ground water. 

A.2.1 TRADITIONAL DRILLING TECHNIQUES 

Traditional drilling techniques include those methods that traditionally have been used to install 
drinking water supply wells. Examples of traditional drilling techniques include hollow stem auger, 
rotary, air percussion, and cable tool or chain tool. They have in common the advantage of being 
capable of installing wells of varying diameters to drinking water well specifications. Each of these 
techniques also allows for visual description of the materials and can allow for easy stratigraphic 
correlation. In general, the equipment required by each of these techniques is readily available. 
Disadvantages of traditional drilling techniques include their expense, time requirements, and waste 
generation. Not only do these techniques produce soil/fluids from the drilling process, frequently, in 
order to properly develop wells by these techniques, a large volume of ground water must be ex¬ 
tracted during a lengthy development. Although the advantages and disadvantages listed above are 
common to most traditional drilling techniques, they are applicable to varying degrees. Furthermore, 
drilling depth and subsurface stratigraphy are important considerations when evaluating the efficacy 
of each of these techniques. 

Hollow-stem auger has been the most widely used traditional drilling technique in environmen¬ 
tal investigations, because it is very effective in the most commonly investigated geologic setting 
encountered during environmental investigations: unconsolidated deposits at shallow depths. Al¬ 
though less common, a chain tool can also be effective under similar geologic conditions. When 
installing wells, a chain tool may require a little more time, but may prove to be less disruptive to the 
formation in the vicinity of the well screen. Both techniques are well suited to collecting continuous 
soil samples using a split-barrel continuous sampling device. This capability is extremely important 
because detailed knowledge of the subsurface can be critical to the successful demonstration of 
remediation by monitored natural attenation. 


A2-5 


At greater depths and in more competent formations, rotary and air hammer techniques are 
frequently used. Rotary techniques are also suited to penetration of cobbly units that may prove 
difficult or impenetrable to a hollow-stem auger or chain tool. With rotary rigs, the fastest drilling 
rates are usually achieved by using drilling fluids such as mud or water; however, these fluids may 
require handling as IDW and may clog the pore space in the vicinity of the well screen. As long as 
air circulation can be maintained in the borehole, an air hammer can be particularly useful in compe¬ 
tent bedrock formations without introducing drilling fluids. 

A.2.2 CONE PENETROMETER TESTING 

CPT is increasingly being used for successful site characterization. CPT is accomplished using 
a cone penetrometer truck, which consists of an instrumented probe that is forced into the ground 
using a hydraulic load frame mounted on a heavy truck, with the weight of the truck providing the 
necessary force. Penetration force is typically supplied by a pair of large hydraulic cylinders bolted 
to the truck frame. In tight soils, push capacity is more often limited by the structural bending 
capacity of the push rods than by the weight of the truck. Cone penetrometers operate well in most 
unconsolidated deposits; however, they may not be able to penetrate and may be damaged by 
cobbles, gravel layers, very stiff clays, and cemented units. 

The penetrometer probe generally consists of a 60-degree conical tip attached to a friction 
sleeve. Inside the probe, two load cells independently measure the vertical resistance against the 
conical tip and the side friction along the sleeve. Each load cell is a cylinder of uniform cross sec¬ 
tion inside the probe which is instrumented with four strain gauges in a full-bridge circuit. Forces 
are sensed by the load cells, and the data are transmitted from the probe assembly via a cable running 
through the push tubes. The analog data are digitized, recorded, and plotted by computer in the 
penetrometer truck. Penetration, dissipation, and resistivity data are used to determine site strati¬ 
graphy. 

The cone penetrometer can be a very effective tool for collecting large quantities of subsurface 
information in a short period of time with virtually no waste generation. A cone penetrometer also 
can be used for installation of ground-water monitoring points, and specially equipped penetrometers 
can be used to screen for mobile and residual fuel hydrocarbon contamination using laser induced 
fluorescence (LIF). Although the equipment is fairly expensive, the overall efficiency can make this 
option relatively inexpensive. 

Most of the disadvantages of CPT are linked to the advantages. For instance, the speed and 
minimal waste associated with CPT are directly related to the process of determining lithology in 
situ; however, this does not allow for visual description of subsurface materials. Isolated soil 
samples can be retrieved for visual description to calibrate the cone penetrometry log, but the proce¬ 
dure cannot be performed frequently (nor continuously) without impairing the efficiency of the 
penetrometer. And while CPT can be very effective at precisely determining changes in lithology on 
the basis of grain size, the lack of a visual description prevents stratigraphic correlation on the basis 
of other parameters, such as color. The U.S. DoD supports a technology development program for 
site characterization using cone penetrometers (the SC APS program). SC APS has developed a 
down-hole CCD camera and light source that can visualize subsurface sediments. 

Monitoring points installed using a cone penetrometer illustrate another advantage that comes 
with disadvantages. CPT allows for rapid placement of discreet ground-water sampling points at a 
precise depth selected on the basis of real-time, detailed, stratigraphic logs. The most effective 
emplacement technique allows for installation of monitoring points of not greater than approximately 
0.5 inch ED. While these points may not require much development or purging, ground-water extrac¬ 
tion for development, purging, and sampling becomes extremely inefficient if the depth to ground 


A2-6 


water is greater than approximately 25 feet. In addition, the monitoring point emplacement tech¬ 
nique typically does not allow for installation of a sand pack, bentonite seal, and grout slurry as may 
be required by regulations. 

A.2.3 HYDRAULIC PERCUSSION SYSTEMS 

A variety of sampling tools can be advanced through unconsolidated soils using relatively 
inexpensive hydraulically powered percussion/probing machines (e.g., Geoprobe®). These sorts of 
systems are frequently mounted on pickup trucks or all-terrain vehicles and, as a result of their small 
size and versatility, can access many locations that larger equipment cannot. 

Hydraulic percussion systems provide for the rapid collection of soil, soil gas, and ground-water 
samples at shallow depths while minimizing the generation of investigation-derived waste materials. 
Specifically undisturbed, continuous soils samples can rapidly be collected for visual observation, 
field analysis, and/or laboratory analysis. In addition, ground-water samples can be collected 
through the probe rods, or ground-water monitoring points can be installed for later sample collec¬ 
tion. Although monitoring points installed by hydraulic percussion systems can vary considerably in 
design and can include sandpacks and seals, monitoring points are typically narrow in diameter. As a 
result, it can be difficult to sample points where the ground-water elevation is greater than 25 feet 
bgs. Furthermore, the narrow diameter may not comply with regulatory standards or future use 
needs. 


A2-7 


SECTION A-3 

SOIL CHARACTERIZATION METHODOLOGIES 

As part of an evaluation of monitored natural attenuation for contaminants in ground water, soil 
characterization factors into development of a site conceptual model, estimation of continuing source 
strength, and modeling of fate and transport. The following sections describe soil sample acquisi¬ 
tion, description, field screening, and laboratory analysis procedures. Samples should be collected in 
accordance with local, State, and Federal requirements. 

A.3.1 SAMPLE ACQUISITION 

Soil samples can be collected using a variety of methods, depending upon the method used to 
advance boreholes. In all cases, the goal is to collect samples to allow lithologic logging and to 
provide useable samples for field screening and for submission to an analytical laboratory. The 
samples should meet the appropriate data quality objectives as identified in the site-specific Quality 
Assurance Plan. 

When using hollow-stem auger or chain tool methods, relatively undisturbed continuous soil 
samples can be collected with split-barrel samplers that are either advanced using a hydraulic ham¬ 
mer or are driven along with the advancing auger. These are well-tested methods that are useful in 
most types of soils except for saturated sands, in which samples tend to liquify and slide out of the 
barrel. Collection of continuous samples allows a more thorough description of site geology, with 
only a slight increase in the time required for drilling. These methods also can be used to collect 
samples in various types of liners, such as acetate or brass sleeves. These sleeves can be cut, capped, 
and shipped with a minimum of effort. When using sleeves, the samples are disturbed less, but 
description of the soils may be hindered if the liners are not clear. Other traditional drilling methods 
(i.e., rotary) do not produce samples that can be used for chemical analysis, and will also make 
geologic interpretation more difficult due to the disturbed nature of the material. 

If CPT or hydraulic percussion methods are used, soil sampled can be collected using a hydrau¬ 
lically driven sampler. When soil samples are collected using a probe-drive sampler, the probe-drive 
sampler serves as both the driving point and the sample collection device and is attached to the 
leading end of the driving rods. To collect a soil sample, the sampler is pushed or driven to the 
desired sampling depth, the drive point is retracted to open the sampling barrel, and the sampler is 
subsequently pushed into the undisturbed soils. The soil cores are retained within brass, stainless 
steel, or clear acetate liners inside the sampling barrel. The probe rods are then retracted, bringing 
the sampling device to the surface. The soil sample can then be extruded from the liners for litho¬ 
logic logging, or the liners can be capped and undisturbed samples submitted to the analytical labora¬ 
tory for testing. 

If a hand auger is used, samples will be slightly disturbed, but still useful for logging purposes. 
Removing soil from the auger bucket may prove difficult where soils are clayey. Below the water 
table, it may be impossible to retain sandy soils in the bucket. Hand driven samplers are similar to 
probe-drive samplers, except that all pushing power is provided manually, 

Following sample acquisition, the coordinates and elevation of all soil sampling locations 
should be surveyed. Horizontal coordinates should be measured to the nearest 0.1 foot relative to an 
established coordinate system, such as state planar. The elevation of the ground surface also should 
be measured to the nearest 0.1 foot relative to USGS mean sea level (msl) data. 

A.3.2 PHYSICAL DESCRIPTION 

Physical characterization of soils should be performed at all sampling locations and a descrip¬ 
tive log prepared for the materials encountered. If using CPT, the descriptive logs should consist of 
continuous computer-generated interpretations supplemented by periodic sensory confirmation and 


A3-8 


description. Otherwise, continuous sampling with interpretation and description is recommended in 
order to precisely identify and isolate changes in lithology. The descriptive log should contain: 

• Sample interval (top and bottom depth); 

• Sample recovery; 

• Presence or absence of contamination; 

• Lithologic description, including relative density, color, major textural constituents, minor 
constituents, porosity, relative moisture content, plasticity of fines, cohesiveness, grain 
size, structure or stratification, relative permeability, and any other significant observa¬ 
tions; and 

• Depths of lithologic contacts and/or significant textural changes measured and recorded to 
the nearest 0.1 foot. 

In addition, representative samples should be photographed, labeled, and stored. Additional site 
characterization features are frequently being added to the list of desirable parameters. Static pore 
pressure and transient pore pressures during penetration with a cone penetrometer are examples. 

A.3.3 FIXED-BASE LABORATORY ANALYSES 

Portions of selected samples should be sent to the fixed-base laboratory for analysis. It is 
desirable to sample and submit a relatively undisturbed sample, if possible. Undisturbed samples are 
typically collected in brass, stainless steel, or clear acetate liners inside of a sampling barrel. Upon 
removal from the barrel, liners are cut to length (if desired) and capped. If the selected drilling 
technique, site conditions, or project requirements do not permit collection of undisturbed soils, 
samples for analysis of volatile constituents should be transferred immediately to an appropriate 
container in such a way as to minimize volatilization during the transfer and headspace in the sample 
container. The analytical protocol to be used for soil sample analysis is presented in Table 2.1. This 
analytical protocol includes the parameters necessary to document the effects of sorption and to 
estimate the magnitude of the continuing source. The protocol document describes each soil analyti¬ 
cal parameter and the use of each analyte in the demonstration of remediation by monitored natural 
attenuation. 

Each laboratory soil sample will be placed in an analyte-appropriate sample container and 
delivered as soon as possible to the analytical laboratory for analysis of total hydrocarbons, aromatic 
hydrocarbons, VOCs, and moisture content using the procedures presented in Table 2.1. In addition, 
at least two samples from locations upgradient, crossgradient, or far downgradient of the contami¬ 
nant source will be analyzed for TOC, and the chemical and geochemical parameters necessary to 
characterize the processes and rates of reaction occurring within the plume. 


A3-9 


SECTION A-4 

GROUND-WATER CHARACTERIZATION METHODOLOGIES 

This section describes the scope of work required to collect ground-water quality samples and 
to perform field analyses to evaluate the demonstration of remediation by monitored natural attenua¬ 
tion. Ground-water sampling should be conducted only by qualified scientists and technicians 
trained in the conduct of well sampling, sampling documentation, and chain-of-custody procedures. 
In addition, sampling personnel should thoroughly review this protocol document and the site- 
specific work plan and quality assurance plan prior to sample acquisition and have a copy of the 
work plan and quality assurance plan available onsite for reference. Samples should be collected in 
accordance with local, State, and Federal requirements. 

A.4.1 GROUND-WATER MONITORING LOCATIONS, DEPTHS, AND SCREENED 
INTERVALS 

Ground-water monitoring locations should be selected on the basis of the preliminary concep¬ 
tual site model and information on the three-dimensional distribution of contaminants. At a mini¬ 
mum, one monitoring location should be placed upgradient from the contaminant plume, one loca¬ 
tion should be placed in the suspected source area, two locations should be placed within the plume, 
and three locations should be placed various distances downgradient and crossgradient from the 
plume. The actual number of monitoring locations could be considerably higher and should be 
related to site conditions and the size of the source. 

It is necessary to collect samples that document the vertical extent of contamination at several 
or at all of the ground-water monitoring locations. This decision is based on the presence of confin¬ 
ing units, the thickness of the aquifer, the type and source of contamination, and suspected variations 
in subsurface transmissivity. The position of well screens should be selected by the field scientist 
after consideration is given to the geometry and hydraulic characteristics of the stratum in which the 
well will be screened. Wells should be screened so that the vertical distribution of contaminants and 
hydraulic gradients can be delineated. Typically the shallowest ground-water monitoring depth is 
chosen to intersect the water table. This allows for the monitoring of LNAPL and seasonal water 
level fluctuations, as well as dissolved contaminant concentrations in the portion of the aquifer 
closest to the typical source. Deeper locations are selected on the basis of contaminant distribution, 
typically above or below suspected confining units or in zones believed to possess higher transmis¬ 
sivity. To ensure well integrity, clustered monitoring wells/monitoring points generally should be 
completed in separate boreholes. 

Screen lengths of not more than 5 feet are recommended to help mitigate the dilution of water 
samples from potential vertical mixing of contaminated and uncontaminated ground water. Screen¬ 
ing a larger area of the saturated zone will result in averaging of contaminant concentrations and 
hydraulic properties. In addition, short screened intervals used in nested pairs give important infor¬ 
mation on the nature of vertical hydraulic gradients in the area. 

A.4.2 TYPES OF GROUND-WATER SAMPLING LOCATIONS 

Ground-water samples for the demonstration of remediation by monitored natural attenuation 
can be collected from monitoring wells, monitoring points, or grab sampling locations. Monitoring 
points and grab locations provide rapid and inexpensive access to shallow ground-water, and yield 
ground-water samples that are appropriate for site characterization and plume definition. Conven¬ 
tional monitoring wells are required for sites with ground-water elevations more than approximately 
25 feet below ground surface. They also are recommended for long-term monitoring (LTM) and 
performance evaluation ground-water sampling, and may be required for regulatory compliance. 

Following installation, the location and elevation of all ground-water monitoring locations 
should be surveyed. Horizontal coordinates should be measured to the nearest 0.1 foot relative to an 


A4-10 


established coordinate system, such as state planar. The elevation of the ground surface also should 
be measured to the nearest 0.1 foot relative to USGS mean sea level (msl) data. Other elevations, 
including the measuring point, should be measured to the nearest 0.01 foot. 

A.4.2.1 Monitoring Weils 

Monitoring wells are commonly installed to evaluate remediation by monitored natural attenua¬ 
tion. As used in this document, monitoring wells are assumed to have, at a minimum, a sand pack, a 
bentonite seal, an annular seal, a surface seal, and an inside diameter of at least 2 inches. Monitoring 
wells are extremely versatile and can be used for ground-water sampling, aquifer testing, product 
recovery systems, long-term monitoring, and performance evaluation monitoring. Although versa¬ 
tile, monitoring wells are relatively expensive to install and create relatively large quantities of waste 
during installation, development, and sampling. Detailed well installation procedures are described 
in the following paragraphs. Of course, local protocols, regulations, type of drill rig, site conditions 
and site-specific data uses should dictate actual well completion details. 

The monitoring well should be installed in a bore hole with a diameter at least 4 inches larger 
than the outside diameter of the well. At a minimum, blank well casing and screen should be con¬ 
structed of Schedule 40 polyvinyl chloride (PVC) with an inside diameter (ID) of 2 inches. Fre¬ 
quently, this diameter must be increased if the well may be used for a pumping test or certain types 
of product or ground-water recovery. The screens should be factory slotted with appropriately sized 
openings (typically 0.010-inch). All well sections should be flush-threaded; glued joints should not 
be used. The casing at each well should be fitted with a threaded bottom plug and a top cap con¬ 
structed of the same type of material as the well casing. The top should be vented to maintain 
ambient atmospheric pressure within the well casing. It is possible that PVC will not be suitable for 
use in wells intended to monitor high concentrations of volatile organic constituents. 

Once the well is in place, sand, bentonite, and grout are used to fill the remaining borehole 
annulus. Appropriately-sized sand must be packed along the entire length of the screen; however, it 
is desirable to limit the vertical distance that the sand pack extends to either side of the screen (i.e., at 
least 6 inches but less than 2 feet) because the added sand pack can increase the portion of the 
aquifer that is effectively screened. A bentonite seal is placed on top of the sand pack. If conditions 
permit, this seal should have a minimum thickness of 2 feet. A cement-bentonite grout is used to fill 
the remainder of the annular space between the bentonite seal and the surface completion. Depend¬ 
ing on site conditions and facility preferences, either flush-mount or stick-up surface completions 
can be used. Site conditions and local, State, and Federal requirements should ultimately dictate 
materials selection and construction details. 

The field scientist should verify and record the boring depth, the lengths of all casing and screen 
sections, and the depth to the top of all well completion materials placed in the annulus between the 
casing and borehole wall. All lengths and depths should be measured to the nearest 0.1 foot. 

A.4.2.2 Monitoring Points 

Where site conditions and the regulatory environment permit, monitoring points are ideal tools 
for rapidly and cost-effectively obtaining site data to evaluate a remediation by monitored natural 
attenuation. Monitoring points can be installed and sampled rapidly while generating a minimal 
volume of waste. Furthermore, some monitoring points cannot be used for ground-water or free 
product level measurements. It is always useful when a site has a reasonable and adequate number 
of monitoring wells. Detailed monitoring point installation procedures are described in the following 
paragraphs. Of course, local protocols, regulations, available equipment, and site conditions should 
dictate actual well completion details. 

In this document, monitoring points are considered temporary or permanent ground-water 
sampling locations that do not meet the specifications of monitoring wells. Typically monitoring 


A4-11 


points are installed in small diameter boreholes using CPT, hydraulic percussion, or manually- 
powered equipment. As a result, monitoring points usually have an ID of less than 2 inches. In 
addition, because of the extremely small to nonexistent annular space between the borehole wall and 
the monitoring point materials, they seldom have a sand pack, bentonite seal, and grout seal, particu¬ 
larly with an annulus of 2 inches. Because these components are missing, ground-water monitoring 
points should be installed only in shallow aquifers where installation of such devices will not result 
in the cross-contamination of adjacent water-bearing strata. 

Like monitoring wells, monitoring points are typically constructed of Schedule 40 PVC casing 
and screen; however, monitoring points also can be constructed from Teflon®-lined tubing attached 
to a stainless steel, wire mesh screen. Because the screens are often installed without a sand pack, a 
slot size of 0.010 inch or smaller should be used. All monitoring point casing and screen sections 
should be flush-threaded; glued joints should not be used. The casing at each monitoring point 
should be fitted with a bottom cap and a top cap constructed of PVC. The top cap should be vented 
to maintain ambient atmospheric pressure within the monitoring point casing. Site conditions and 
local, State, and Federal requirements should ultimately dictate materials selection and construction 
details. 

The field hydrogeologist should verify and record the total depth of the monitoring point, the 
lengths of all casing and screen sections, and the depth to the top of all monitoring point completion 
materials. All lengths and depths should be measured to the nearest 0.1 foot. 

A.4.2.3 Grab Sampling 

Ground-water grab samples are temporally and spatially discrete samples collected from bore¬ 
holes that are abandoned upon completion of sampling. In highly transmissive aquifers, the collec¬ 
tion of grab samples can provide a rapid, cost-effective alternative to the use of monitoring points. 
Like monitoring points, collection of grab samples generates minimal waste; however, they are not 
appropriate for aquifer testing, remediation systems, or long-term monitoring. Furthermore, because 
the locations are abandoned upon completion of sampling, analytical results cannot be confirmed, 
and ground-water levels at all locations cannot be collected over the space of a few hours for use in 
the development of ground-water flow maps. In addition, if the aquifer is not particularly transmis¬ 
sive, sample collection can require hours resulting in inefficient equipment utilization. For these 
reasons, installation and sampling of monitoring points typically is recommended where feasible. 
Several of the more common instruments used to collect ground-water grab samples include the 
HydroPunch®, Geoprobe®, cone penetrometer, or hand-driven points. An optimal site characteriza¬ 
tion approach often involves use of grab samples acquired by push technologies such as the 
HydroPunch®, Geoprobe®, cone penetrometer, or hand-driven points for a rapid, three-dimensional 
characterization of the site, then using that information to select locations and screened intervals for 
permanent monitoring points. 

A.4.3 MEASUREMENT OF STATIC FLUID LEVELS 

A.4.3.1 Water Level and Total Depth Measurements 

Prior to purging or developing any water from a ground-water sampling location, the static 
water level should be measured. At all locations of sufficient diameter, an electric water level probe 
should be used to measure the depth to ground water below the datum to the nearest 0.01 foot. Small 
diameter probes are commercially available for measurement of water levels in monitoring points 
and through Geoprobe®, HydroPunch®, and CPT pushrods. After measuring the static water level, 
the water level probe should be slowly lowered to the bottom of the well, and the total well depth 
should be measured to the nearest 0.01 foot. If measuring from the ground surface, an accuracy 
better than 0.1 foot is probably not practical. Based on these measurements the volume of water to 
be developed or purged from the location can be calculated. If mobile LNAPL is encountered, the 


A4-12 


LNAPL thickness should be determined, and attempts should be made to sample both the ground 
water below the LNAPL layer as well as the LNAPL. 

If a sufficiently narrow water level probe is unavailable, hollow, high-density polyethylene 
(HDPE) tubing connected to a manometer can be used to determine depth to ground water. The 
manometer will indicate when ground water is reached as the HDPE tubing is inserted into the 
monitoring location. The HDPE attached to the manometer will then be marked at the level of the 
ground surface and removed. The depth to water will be determined by placing a tape measure next 
to the HDPE tubing and measuring the length from the base of the tubing to the ground level mark to 
the nearest 0.01 foot, if possible. 

A.4.3.2 Mobile LNAPL Thickness Measurements 

At sites where phase-separated hydrocarbons are present in the ground-water system, it is 
important to accurately measure the thickness of floating hydrocarbons. Accurate measurement of 
hydrocarbon thickness allows for estimation of the amount and distribution of the hydrocarbon and 
correction of measured ground-water elevations. There are three methods that can be used to deter¬ 
mine the thickness of mobile LNAPL in a well, including use of an interface probe, a bailer, or tape 
and paste. Interface probes generally operate on either light refraction sensors or density float 
switches to detect hydrocarbons and the hydrocarbon/water interface. The depth to mobile LNAPL 
and depth to water should be measured to the nearest 0.01 foot. The thickness of phase-separated 
hydrocarbons should also be measured to the nearest 0.01 foot. Three consecutive measurements 
should be made to ensure the accuracy of the measuring instrument. A clear bailer can be slowly 
lowered into the well until it intersects the fluid but is not totally immersed. The bailer is then 
retrieved, and the floating LNAPL can be visually observed and measured with an engineer’s tape. 
The third method for measurement of floating hydrocarbon thickness is hydrocarbon paste and an 
engineer’s tape. The paste, when applied to the tape, changes color when it intersects the hydrocar¬ 
bon and the hydrocarbon/water interface. Measurements of the mobile LNAPL thickness can be 
made directly from the engineer’s tape. It is extremely important to remember to thoroughly decon¬ 
taminate all equipment between well measurement events to prevent cross-contamination of wells. 
Equipment blanks, part of the Quality Assurance Program, will confirm the suitability of the decon¬ 
tamination activities. 

Measurements of mobile LNAPL thickness made in monitoring wells provide only an estimate 
of the actual thickness of NAPL at that location. Actual mobile and residual LNAPL thicknesses can 
only be obtained from continuous soil cores. Correcting apparent mobile LNAPL thickness as 
measured in monitoring wells to true thickness is discussed in Appendix C. 

A.4.3.3 Mobile DNAPL Thickness Measurements 

DNAPL thickness in wells cannot be used to estimate actual DNAPL quantities on a site. 

A.4.3 GROUND-WATER EXTRACTION 

Varied equipment and methods are available for the extraction of ground water. The approach 
is determined on the basis of application (development, purging, or sampling), hydrogeologic condi¬ 
tions, monitoring location dimensions, and regulatory requirements. 

Ground water produced during extraction activities must be handled in a manner consistent with 
the investigation-derived waste (IDW) plan for the site. The method of handling and disposal will 
depend on location and type of source, site contaminants, degree of contamination (e.g., free product, 
odor, air monitoring measurements), and applicable local, State, and Federal regulations. 

A.4.3.1 Methods 

Portable ground-water extraction devices from three generic classifications are commonly used 
for investigations of monitored natural attenuation: grab, suction lift, and positive displacement. 


A4-13 


The selection of the type of device(s) for the investigation is based on type of activity, well/point 
dimensions, and hydrogeologic conditions. 

Bailers are common grab sampling devices. Disposable bailers can be used to avoid decontami¬ 
nation expenses and potential cross-contamination problems. Drawbacks for bailers include agita¬ 
tion/aeration of the ground water and the inability to maintain a steady, non-turbulent flow required 
to establish a true flow-through cell. Aeration also can be an issue during transfer of the sample 
from the bailer to the sample container. As a result of aeration, and because a true flow-through cell 
cannot be established, accurate dissolved oxygen and ORP measurements can be difficult to obtain. 

The suction lift technology is best represented in environmental investigations by the peristaltic 
pump. A peristaltic pump extracts water using a vacuum created by cyclically advancing a sealed 
compression along flexible tubing. This pumping technique means that extracted water contacts 
nothing other than tubing that can be easily replaced between sampling locations. This reduces the 
possibility of cross-contamination. Furthermore, peristaltic pumps can be used to extract minimally- 
disturbed ground water from any size monitoring location at variable low-flow rates. Because of 
these features, representative samples are simple to collect, and reliable flow-through cells are 
simple to establish. The biggest drawback with a peristaltic pump is the maximum achievable 
pumping depth which is equivalent to the height of water column that can be supported by a perfect 
vacuum. This effectively limits the use of a peristaltic pump to monitoring locations with ground- 
water depths of less than approximately 25 feet. Also, off-gasing can occur in the tubing as a result 
of the reduced pressures and high-rate of cyclical loading. If bubbles are observed in the tubing 
during purging or sampling, the flow rate of the peristaltic pump must be slowed. If bubbles are still 
apparent, the tubing should be checked for holes and replaced. The final potential disadvantage with 
a peristaltic pump is the low flow rate. Although advantageous for sampling, this can be inappropri¬ 
ate during purging or development at locations with large extraction volumes. Puls and Barcelona 
(1996) show that the use of peristaltic pumps does not compromise sample integrity as long as no 
bubbles form during sampling. If the ground water is saturated with methane or carbon dioxide, it is 
practically impossible to collect samples without a gas headspace. Pankow (1986) gives advice on 
how to correct for this problem. 

Positive displacement pumps, also called submersible pumps, include, for example, bladder 
pumps, Keck®, Grundfos Redi-Flo II®, Bennett® and Enviro-Tech Purger ES® pumps. Each of these 
pumps operates downhole at depths of up to a few hundred feet and rates of up to several gallons per 
minute. Therefore, submersible pumps are particularly useful for applications requiring the extrac¬ 
tion of large volumes of water or for the extraction of ground water from depths in excess of 25 feet. 
Because the pumps operate downhole, they require appropriately-sized wells. At a minimum, an 
inside well diameter of at least 1.5 inches typically is required; however, much larger well diameters 
can be required depending on the selected pump type, extraction depth, and extraction rate. Because 
typical submersible pump design results in contact between the ground water and internal as well as 
external surfaces of the pump, rigorous decontamination and quality assurance procedures must be 
implemented to avoid cross-contamination if a pump that is not dedicated to the well is used for 
sampling. 

A.4.3.2 Development 

Monitoring wells and points should be developed prior to sampling to remove fine sediments 
from the portion of the formation adjacent to the screen. Development is not required for grab 
sampling locations. Because development is intended to enhance ground-water production and 
quality through the removal of fine sediments in the immediate vicinity of the screen, high flow rates 
and downhole turbulence are beneficial. This is particularly true for monitoring wells because of the 
formation disturbance usually associated with installation. Development can be accomplished using 


A4-14 


any of the methods discussed in Section A.4.3.1 with selection dependent on well/point dimensions, 
well/point installation procedures, and hydrogeologic conditions. 

Development is accomplished through the removal of water from the well/point in combination 
with screen/sand pack cleansing through agitation of the downhole ground water. The “agitation” is 
typically provided by pumping at a high flow rate; surging with the pump, a surge block, or a bailer; 
and/or pumping along the entire length of the screen. As a rule, the more “agitation” that can be 
provided, the “better” the development. Typically during development, ground water is extracted 
until dissolved oxygen, pH, temperature, specific conductivity, and water clarity (turbidity) stabilize. 
Monitoring well/point development should occur a minimum of 24 hours prior to sampling. Devel¬ 
opment water must be handled in accordance with the site IDW plan. 

It is important to maintain a record of development for each location. The development record 
should include the following information, at a minimum: 

• Monitoring point/well number; 

• Date and time of development; 

• Development method; 

• Monitoring point/well depth; 

• Volume of water produced; 

• Description of water produced; 

• Post-development water level and monitoring point/well depth; and 

• Field analytical measurements, including pH, temperature, and specific conductivity. 

A.4.3.3 Purging 

Purging consists of the evacuation of water from the monitoring location prior to sampling, so 
that “fresh” formation water will enter the monitoring location and be available for sampling. Be¬ 
cause sampling can occur immediately upon completion of purging, it is best to limit ground-water 
agitation, and consequently, aeration of the ground water and volatilization of contaminants. Two 
sources for agitation include the purging device and the cascading of water down the screen as the 
water level in the well drops. To avoid agitation, a low-disturbance device such as a peristaltic pump 
or bladder pump is recommended for purging, while equipment such as bailers should be avoided. 

To avoid aeration, wells or points that were initially screened below the water table should be 
pumped at a rate which prevents lowering of the water table to below the top of the screen, and if 
practical, wells or points screened across the water table should be pumped at a rate that lowers the 
total height of the water column no more than 10 percent of the screened interval. Purging should 
follow the recommendations of Puls and Barcelona (1996). 

Typically, the volume of water contained within the monitoring well/point casing is used to 
estimate the amount of ground water that should be removed during the purge. As a general rule, 
three times the calculated volume should be removed from the well/monitoring point; however, this 
can be reduced to between 1 and 3 volumes for low-producing wells and wells with a very large 
water column, but a very short screened interval. Purging should continue until parameters such as 
pH, temperature, specific conductance, dissolved oxygen, and ORP stabilize. Sampling should occur 
as soon after purging as practical, and definitely within 24 hours. Purge waters must be handled in 
accordance with the site IDW plan. 

If a monitoring well/monitoring point is evacuated to a dry state during purging, the monitoring 
well/monitoring point should be allowed to recharge, and the sample should be collected as soon as 
sufficient water is present in the monitoring well or monitoring point to obtain the necessary sample 
quantity. Sample compositing or sampling over a lengthy period by accumulating small volumes of 
water at different times to obtain a sample of sufficient volume should be avoided. 


A4-15 



It is important to record purge information as a part of the sampling record for each location. At 
a minimum, the following information pertaining to the purge should be recorded: 

• Monitoring point/well number; 

• Date and time of purge; 

• Purge method; 

• Monitoring point/well depth; 

• Volume of water produced; 

• Description of water produced; 

• Post-purge water level; and 

• Field analytical measurements, including pH, temperature, specific conductivity, dissolved 
oxygen concentration, and ORP; 

• Thickness of LNAPL, if present, in the point/well prior to purging; 

• Volume of LNAPL removed during purging. 

A.4.3.4 Sampling 

Sampling should occur immediately after purging. If well yield is less than 1/10 of a liter per 
minute, sample according to the guidance provided by Puls and Barcelona (1996). The object of 
sampling is the collection of representative ground-water samples. This means that impact to the 
sample as a result of turbulence, contact with equipment, or a change in conditions must be mini¬ 
mized. The use of a peristaltic pump with dedicated HDPE tubing is recommended for monitoring 
locations where the depth to water is less than 25 feet because the peristaltic pump is capable of 
providing a steady, low-flow, stream of ground water which has contacted only dedicated tubing. In 
addition, conditions are relatively unchanged, so long as care is taken to ensure that the pumping 
suction does not cause the ground water to boil as a result of the reduced pressure. Where the depth 
to ground water is greater than 25 feet, a dedicated positive displacement pump, when available, is 
best. Because of the decontamination difficulties and the resulting potential for cross-contamination 
associated with most positive displacement pumps, sampling through these pumps is not recom¬ 
mended unless the pumps are dedicated. A bailer should be used only if it is the only means of 
obtaining a sample. 

An overflow cell, such as the one pictured on Figure A.4.1, or a flow-through cell as pictured in 
Figure A.4.2, should be used for the measurement of well-head parameters, including pH, tempera¬ 
ture, specific conductance, dissolved oxygen, qnd ORP. When using a pump to purge or sample, the . 
pump intake tubing should be positioned near the bottom of the cell. If using a bailer, the water 
should be drained from the bottom of the bailer through tubing into the cell. In either case, the 
tubing should be immersed alongside the dissolved oxygen probe beneath the water level in the cell. 
This will minimize aeration and keep water flowing past the dissolved oxygen probe’s sampling 
membrane. The probes for the other parameters are less sensitive to positioning within the flow¬ 
through cell. 

Samples should be collected directly from the pump discharge tube or bailer into a sample 
container of appropriate size, style, and preservation for the desired analysis. Water should be 
directed down the inner walls of the sample bottle to minimize aeration of the sample. All samples 
to be analyzed for volatile constituents (e.g., SW8010, SW8020, SW8240, SW8260, and TPH-g) or 
dissolved gases (e.g., methane, ethane, and ethene) must be filled and sealed so that no air space 
remains in the container. Sample handling procedures are further described in Section A.6. 


A4-16 




Figure A.4.1 Overflow cell to prevent alteration of geochemical properties of ground water by exposure to 
the atmosphere. 


* 



Figure A. 4.2. Flow-through cell to prevent alteration of geochemical properties of ground water by 
exposure to the atmosphere. 


A.4.4 GROUND-WATER ANALYTICAL PROCEDURES 

In order to demonstrate the efficacy of monitored natural attenuation, field and laboratory 
analyses should be performed on all ground-water samples using the analytical procedures listed in 
Table 2.1. As a result of analyte properties and available detection equipment, analyses can be 
performed at the sampling location, a portable field laboratory, or a fixed-base laboratory. The 
dissolved hydrogen analysis is unique in that it requires a combination of well-head and field labora¬ 
tory procedures that are somewhat different from other field methods; therefore, it is presented in a 
separate subsection. Several of the analytes or parameters can be measured in more than one man¬ 
ner; consequently, the methods provided in this section should not be considered absolute. Rather, 
these methods have been proven to provide reliable information. The site-specific data quality needs 


A4-17 




























of each project will be determined during the Data Quality Objective Process and documented in the 
Quality Assurance Plan. 

In order to obtain accurate and defensible data, it is critical that quality assurance procedures are 
followed for all analyses. These procedures generally fall into the following categories: 

• Collection and handling of samples; 

• Calibration of direct read meters, chromatographs, colorimeters, and field instruments per 
manufacturer’s instructions; 

• Decontamination of equipment and containers; and 

• Confirmation of results through analysis of blanks, duplicates, and other quality control 
samples. 

Actual procedures are equipment and analysis specific, and must be developed accordingly. 

A.4.4.1 Standard Well-Head Analyses 

Standard well-head analyses include pH, conductivity, temperature, dissolved oxygen, and ORP 
because these parameters can be measured with a direct-reading meter. This allows all of these 
parameters to be used as indicators for ground-water stability during development and purging 
activities. In addition, dissolved oxygen and ORP can be used to provide real time data on the 
location of the contaminant plume, especially in areas undergoing anaerobic biodegradation. Tem¬ 
perature, dissolved oxygen, and ORP must be measured at the well head in unfiltered, unpreserved, 
“fresh” water because these parameters can change significantly within a short time following 
sample acquisition. Section 2.3.2 of the protocol document describes each analysis and its use in the 
demonstration of monitored natural attenuation. 

It is critical that samples collected for well-head analyses are disturbed and aerated as little as 
possible; therefore, the use of a flow-through cell, as described in Section A.4.3 and illustrated on 
Figure A.4.1, is recommended. Where this is not possible, measurements can be made in a clean 
glass container separate from those intended for laboratory analysis. Where ground-water extraction 
disturbs the sample, downhole probes can be used for dissolved oxygen analyses, but such probes 
must be thoroughly decontaminated between wells. In some cases, decontamination procedures can 
be harmful to the dissolved oxygen probe, and inadequate decontamination can create potential 
cross-contamination problems if performed prior to sample collection for the other analytes. After 
sample acquisition, the downhole ground water may be too disturbed to collect an accurate downhole 
DO measurement. 

A.4.4.2 Dissolved Hydrogen Analysis 

As described in Section 2.3.2.9, dissolved hydrogen (H 2 ) concentrations can be an indicator of 
microbially mediated redox processes in ground-water systems. Determination of H 2 concentrations 
is a two-step process in the field: sampling at the well head and analysis with a reducing gas detector. 

Hydrogen is highly volatile, and this chemical property can be used to measure H 2 concentra¬ 
tions in ground water. The principle is to continuously pump ground water through a gas-sampling 
bulb containing a nitrogen or air “bubble” so that the H 2 can partition between the gas and liquid 
phases until the concentration of H 2 in the bubble comes into equilibrium with concentration of H 2 in 
the ground water. The bubble is then analyzed for H 2 and the concentration of H 2 in the ground 
water is calculated using the Ideal Gas Law and Henry’s Law. This method is referred to as the 
“bubble strip” method (Chapelle et al., 1995,1997), because the bubble “strips” H 2 out of the water. 

A.4.4.2.1 Sampling Method 

The following procedures are recommended for the collection of a sample for analysis by the 
“bubble strip” method: 

1. Place the intake hose of a peristaltic pump, a Bennett positive displacement pump, or a blad¬ 
der pump into the sampling well at the depth of the screened interval. 


A4-18 


Do not sample for H 2 with electrical submersible pumps because they may produce hydrogen. 

Do not sample for H 2 from wells with metal screens or casings because they may produce hydrogen 
and interfere with measurements. 

2. Attach a glass, 250-ml gas-sampling bulb (Figure A.4.3) to the outflow end of the tube. 

3. Turn on the pump and adjust the flow rate to between 400 and 700 mL/min. 

4. Briefly hold the outlet end of the sampling bulb in the upright position to remove any gas 
bubbles from the bulb. 

5. Place the bulb in a horizontal position and inject 20 mL of hydrogen-free N 2 gas through the 
septum (Figure A.4.3). 

6. Allow the N, bubble to come into equilibrium with the flowing ground water for 30 minutes. 
This equilibration process takes approximately 20 minutes. 

7. Remove 3-5 mL of the gas bubble using a 10 mL glass syringe with attached mini-inert valve. 

8. Close the valve to seal the sample. 

9. Wait an additional 5 minutes and repeat steps 7 and 8. 

10. Analyze both samples on the hydrogen detector, as described in Section A.4.4.2.2. 

Resample the well if the H 2 concentrations of the duplicate samples do not agree within 10 percent. 


✓ 



Figure A.4.3 Schematic showing the “bubble strip ” method for measuring dissolved hydrogen 
concentrations in ground water. 


A.4.4.2.2 Analytical Method 

Concentrations of H 2 in the nitrogen bubble are determined by gas chromatography (GC) with 
reduction gas detection (Trace Analytical, Menlo Park, CA). To perform this analysis, a gaseous 
sample is injected into the stream of a carrier gas such as N 2 . The sample is transported by the 
carrier through a separation column where the components of the sample are separated on the basis 
of variations in their transport efficiency through the column matrix. The column is packed with 
CarboSieve II which separates chemical species primarily on the basis of molecular size. The sepa¬ 
rated components elute from the column and pass through a heated bed of HgO where the reduced 
gases (primarily H 2 and CO) are oxidized and Hg vapor is released. The concentration of Hg vapor 
released is directly proportional to the concentration of reduced gases present in the sample and is 


A4-19 



detected by means of an ultraviolet photometer. Because chlorinated solvents can destroy the HgO 
bed, the column is backflushed immediately after the H 2 peak is quantified. 

The concentration of H 2 dissolved in the ground water can be calculated from the equilibrated 
concentration in the nitrogen gas bubble as follows: 

1) Prepare a calibration curve for H 2 using a 100 ppm Scotty II standard gas mixture. The cali¬ 
bration curve should range from 0.1 to 10.0 pL/L (ppm). 

2) Analyze the gas sample taken from the gas-sampling bulb, obtaining results (C B ) in units of 
pL/L (ppm) in the gas phase. 

3) Calculate the aqueous concentration of H 2 (C m in nanomoles per liter (nM)) in equilibrium 
with the equilibrated bubble gas (C B , pL/L (ppm)) sample using the conversion factor: 

C^ = 0.81C 5 eq. A.4.1 

This conversion factor is derived from the Ideal Gas Law and Henry’s Law as follows: 

PV=nRT (Ideal Gas Law) eq. A.4.2 

Rearrange to give: 


n _ P 
V~ RT 


eq. A.4.3 


Where: 

n = the quantity of gas in moles 
V = the volume the gas occupies in Liters 
P = the partial pressure of the gas in atm 
T = the temperature in °K 

R = the gas constant (R = 0.08205 atm L mole 1 °K 1 ) 

Thus the concentration of a pure gas at atmospheric pressure and room temperature is 
40.9mmoles/L. 

For a 1.0 ppm calibration standard (i.e., 1.0 pL/L), the H 2 concentration in molar units would be: 

(40.9mmoles / L H )(\0 6 L h / L gas )(\0 6 nmoles / mmoles) = 40.9nmoles/ L gas eq. A.4.4 

The dissolved H 2 concentration in the aqueous phase is given by Henry’s Law: 


c - c * 


H 


H, 


(40.9 nmoles L x ppm l ) 


.-I- 


Conversion factor = 


50.4 


-0.81 


eq. A.4.5 

t 

eq. A.4.6 


Where: 

C = the dissolved H^ concentration in nmoles/L 

C h = the equilibrated bubble H 2 concentration in nmoles/L 

H H2 = the dimensionless Henry’s Law coefficient for the distribution of H 2 between the 
gaseous and dissolved phases (H H2 = 50.4). 

4) Identify the predominant terminal electron accepting process for the water sample using the 
characteristic ranges presented in Table 2.5. 

A.4.4.3 Field Analytical Laboratory Analyses 

The field analytical laboratory analyses to be used for ground-water samples are presented in 
Table 2.1. These analyses include parameters that are time-sensitive or can be performed accurately, 
easily, and inexpensively on site. In addition, results obtained from field laboratory analyses provide 
real-time data on the location of the contaminant plume, especially in areas undergoing anaerobic 
biodegradation. This real-time data can be used to guide the investigation of monitored natural 


A4-20 





attenuation at sites with limited or ambiguous hydrogeologic and plume information. Section 2.3.2 
of the protocol document describes each analysis and its use in the demonstration of monitored 
natural attenuation. 

In preparation for field laboratory analysis, all glassware or plasticware used in the analyses 
must be cleaned thoroughly by washing with a solution of laboratory-grade, phosphate-free detergent 
(such as Alconox®) and water, and rinsing with deionized water and ethanol to prevent interference 
or cross-contamination between measurements. If concentrations of an analyte are above the range 
detectable by the titrimetric method, the analysis should be repeated by diluting the ground-water 
sample with double-distilled water until the analyte concentration falls to a level within the range of 
the method. All rinseate and sample reagents accumulated during ground-water analysis must be 
handled appropriately, including collection, labeling, storage, and disposal. 

Carbon dioxide (C0 2 ) is a byproduct of naturally occurring aerobic and anaerobic biodegrada¬ 
tion processes that occur in ground water. Carbon dioxide concentrations in ground water can be 
measured in the field by titrimetric analysis using CHEMetrics® Method 4500 (0 to 250 mg/L as 
C0 2 ), or similar. 

An increase in the alkalinity of ground water above background may be produced when carbon 
dioxide produced by biological activity reacts with carbonate minerals in the aquifer matrix material. 
Alkalinity of the ground-water sample will be measured in the field by titrimetric analysis using 
U.S. EPA-approved Hach® Method 8221 (0 to 5,000 mg/L as calcium carbonate), or similar. 

Nitrate-nitrogen concentrations are of interest because nitrate can act as an electron acceptor 
during hydrocarbon biodegradation under anaerobic soil or ground-water conditions. Nitrate-nitro¬ 
gen is also a potential nitrogen source for hydrocarbon-degrading bacteria biomass formation. Ni¬ 
trite-nitrogen is an intermediate byproduct in both ammonia nitrification and in nitrate reduction in 
anaerobic environments. Nitrate- and nitrite-nitrogen concentrations in ground water can be mea¬ 
sured in the field by colorimetric analysis using a portable colorimeter (such as the Hach® DR/700). 
Nitrate concentrations in ground-water samples can be analyzed after preparation with Hach® 

Method 8039 (0 to 30.0 mg/L nitrate), or similar. Nitrite concentrations in ground-water samples can 
be analyzed after preparation with U.S. EPA-approved Hach® Method 8507 (0 to 0.35 mg/L nitrite), 
or similar. 

Sulfate in ground water is a potential electron acceptor for fuel-hydrocarbon biodegradation in 
anaerobic environments, and sulfide is produced by biological sulfate reduction. Sulfate and sulfide 
concentrations can be measured by colorimetric analysis with a portable colorimeter (such as the 
Hach® DR/700) after appropriate sample preparation. U.S. EPA-approved Hach® Methods 8051 (0 
to 70.0 mg/L sulfate) and 8131 (0.60 mg/L sulfide) (or similar) can be used to prepare samples and 
analyze sulfate and sulfide concentrations, respectively. 

Iron IE is an electron acceptor for biological metabolism under anaerobic conditions. Iron III is 
the substrate for biological iron reduction; Iron II is the product. Iron concentrations can be mea¬ 
sured in the field by colorimetric analysis with a portable colorimeter (such as a Hach® DR/700) after 
appropriate sample preparation. Hach® Method 8008 for total soluble iron (0 to 3.0 mg/L ferric + 
ferrous iron) and Hach® Method 8146 for ferrous iron (0 to 3.0 mg/L) (or similar) can be used to 
prepare and quantitate the samples. Ferric iron is quantitated by subtracting ferrous iron levels from 
total iron levels. 

Manganese is a potential electron acceptor under anaerobic environments. Manganese concen¬ 
trations can be quantitated in the field using colorimetric analysis with a portable colorimeter (such 
as a Hach® DR/700). U.S. EPA-approved Hach® Method 8034 (0 to 20.0 mg/L), or similar, can be 
used to prepare the samples for quantitation of manganese concentrations. 


A4-21 


A.4.4.4 Fixed-Base Laboratory Analyses 

The fixed-base laboratory analyses to be used for ground-water samples are presented in 
Table 2.1. These analyses include the parameters that cannot be easily or accurately performed in the 
field, but are necessary to document monitored natural attenuation of fuel hydrocarbons and chlori¬ 
nated solvents in ground water. Section 2.3.2 of the protocol document describes each analysis and 
its use in the demonstration of monitored natural attenution. 

Prior to sampling, arrangements should be made with the analytical laboratory (or other sup¬ 
plier) to provide a sufficient number of appropriate sample containers for the samples (including 
quality control samples) to be collected. All containers, preservatives, and shipping requirements 
should be consistent with the analytical protocol. For samples requiring chemical preservation, 
preservatives are best added to containers by the laboratory (or other supplier) prior to shipping. 
Sample handling is discussed in Section A.6. 


A4-22 


SECTION A-5 

SURFACE WATER AND SEDIMENT CHARACTERIZATION 

METHODOLOGIES 

At sites where surface water bodies are affected (or potentially affected) by contamination, 
surface water and sediment sample collection and analysis may be required as a component of the 
remediation by monitored natural attenuation demonstration. 

A.5.1 SURFACE WATER SAMPLE COLLECTION 

Surface water can be collected with a peristaltic pump using exactly the same equipment and 
procedures to collect water from a well. The sampling tube can be introduced into the water from a 
barge or boat, or from a dock. The depth to the sediment should be sounded, then the tube intro¬ 
duced to a level a very few inches above the sediment layer. A weight can be used to keep the tube 
straight. Alternately, !4 inch PVC pipe can be inserted to the correct depth, then sampled with a tube 
just as if it were a well. 

Many plumes discharge at some distance away from the shoreline of lakes or large rivers. 
Samples should be taken at locations where the elevation of the sediment-to-water interface corre¬ 
sponds to the elevation of the contaminant plume in the aquifer. Many plumes are driven down into 
aquifers by recharge. Conversely, the flow path bends sharply up underneath a gaining stream at the 
point of discharge. Water just above the sediment in the center of a stream or small river should be 
sampled. If possible, the stage of a stream or river at a gauging station near the point of sampling 
should be determined to estimate the discharge of the stream or river at the time of sampling. Losing 
streams or rivers should not be sampled at high stage when they are losing water because groundwa¬ 
ter plumes would be pushed away from the sediment interface. To ensure that the stream is not 
losing, the elevation of standing water in monitoring wells near the river should be higher than the 
stage of the river or stream at the time of sampling. The same considerations apply to tidal environ¬ 
ments or areas with wind seiches on large bodies of water. Surface water should be sampled when 
the tide is out, or the wind is blowing off-shore. Additionally, contaminant plumes may be deflected 
strongly downstream by flow occurring within the saturated material surrounding the surface water 
channel. This is particularly true when the hydraulic conductivity of the stream sediments is much 
greater than the hydraulic conductivity of the surrounding material that supplies ground water to the 
stream. A great deal of thought as to when and where to sample is necessary to yield meaningful 
results. 

A.5.2 SEDIMENT SAMPLE COLLECTION 

Sediment samples below the water surface can be collected using a core barrel. The core barrel 
can be hand driven to the desired depth from a boat, then pulled back up using a mechanical jack 
after sampling is finished. An alternative technique is to place open-end, two-inch diameter PVC 
tubing to a desired depth, then insert flexible tubing and collect the sediment as a slurry into a suc¬ 
tion flask connected to a peristaltic pump. 


A5-23 


SECTION A-6 
SAMPLE HANDLING 

This section describes the handling of soil and ground-water samples from the time of sampling 
until the samples arrive at the laboratory. 

A.6.1 SAMPLE PRESERVATION, CONTAINERS, AND LABELS 

Sample containers and appropriate container lids must be purchased or provided by the analyti¬ 
cal laboratory. Any required chemical preservatives can be added to the sample containers by the 
analytical laboratory prior to shipping the containers to the site or alternatively, at the time of sam¬ 
pling. The sample containers should be filled and tightly sealed in accordance with accepted proce¬ 
dures for the sample matrix and the type of analysis to be conducted. The sample label should be 
firmly attached to the container side, and the following information legibly and indelibly written on 
the label: 

• Facility name; 

• Sample identification; 

• Sample type (groundwater, surface water, etc.); 

• Sampling date; 

• Sampling time; 

• Preservatives added; and 

• Sample collector’s initials. 

A.6.2 SAMPLE SHIPMENT 

After the samples are sealed and labeled, they should be packaged for transport to the analytical 
laboratory. The packaged samples should be delivered to the analytical laboratory shortly after 
sample acquisition using an overnight delivery service. The following packaging and labeling 
procedures are to be followed: 

• Abide by all U.S. Department of Transportation (DOT) shipping regulations; 

• Package samples so that they will not leak, spill, or vaporize from their containers; 

• Place samples in a cooler containing ice to maintain a shipping temperature of approxi¬ 
mately 4 degrees centigrade (°C), if required by the requested analyses; 

• Include a properly completed chain-of-custody form, as described in the following subsec¬ 
tion; and 

• Label shipping container with 

- Sample collector’s name, address, and telephone number; 

- Laboratory’s name, address, and telephone number; 

- Description of sample; 

- Quantity of sample; and 

- Date of shipment. 

A.6.3 CHAIN-OF-CUSTODY CONTROL 

After the samples are collected, chain-of-custody procedures must be followed to establish a 
written record of sample handling and movement between the sampling site and the analytical 
laboratory. Each shipping container should include a chain-of-custody form completed in triplicate 
by the sampling personnel. One copy of this form should be kept by the sampling contractor after 
sample delivery to the analytical laboratory; the other two copies should be retained at the labora¬ 
tory. One of the laboratory copies will become a part of the permanent record for the sample and 
will be returned with the sample analytical results. The chain-of-custody form should contain the 
following information: 


A6-24 



• Unique sample identification number; 

• Sample collector’s printed name and signature; 

• Date and time of collection; 

• Sample location; 

• Sample matrix; 

• Sample size and container; 

• Chemical preservatives added; 

• Analyses requested; 

• Signatures of individuals involved in the chain of possession; and 

• Inclusive dates of possession. 

The chain-of-custody documentation should be placed inside the shipping container so that it 
will be immediately apparent to the laboratory personnel receiving the container, but cannot be 
damaged or lost during transport. The shipping container is to be sealed so that it will be obvious if 
the seal has been tampered with or broken. 

A.6.4 SAMPLING RECORDS 

In order to provide complete documentation of the sampling event, detailed records must be 
maintained by the field scientist. At a minimum, these records must include the following informa¬ 
tion: 

• Sample location (facility name); 

• Sample identification; 

• Sample location map or detailed sketch; 

• Date and time of sampling; 

• Sampling method; 

• Field observations of 

- Sample appearance, 

- Sample odor; 

• Weather conditions; 

• Water level prior to purging (ground-water samples); 

• Total well depth (ground-water samples); 

• Purge volume (ground-water samples); 

• Water level after purging (ground-water samples); 

• Well condition (ground-water samples); 

• Sample depth; 

• Sampler’s identification; 

• Field measurements such as pH, temperature, specific conductivity, dissolved oxygen 
concentration, and redox potential (ground-water samples); and 

• Any other relevant information. 


A6-25 





SECTION A-7 

AQUIFER CHARACTERIZATION METHODOLOGIES 

Adequate characterization of the ground-water flow and contaminant transport system is an 
important component of the monitored natural attenuation demonstration. The following sections 
describe methodologies that are recommended to characterize the hydrogeologic system. 

A.7.1 HYDRAULIC CONDUCTIVITY 

Hydraulic conductivity is a measure of an aquifer’s capacity to transmit water and governs 
ground-water flow and contaminant transport in the subsurface. Methods for determining hydraulic 
conductivity in the field can include slug tests, pumping tests, and downhole flowmeter measure¬ 
ments. Hydraulic conductivity can also be measured during penetration with a cone penetrometer by 
measuring the transient pressure excursions in the pore water in front of the cone using a cone 
equipped with a pressure transducer in contact with the pore water. The method selected for a given 
site will depend on the dimensions, locations, and screened intervals of site wells and monitoring 
points; site stratigraphy; equipment availability; budget; and waste handling requirements. 

A.7.1.1 Pump Tests 

A pumping test involves pumping one well at a constant rate for a specified length of time and 
collecting periodic water level measurements in both the pumped well and nearby observation wells 
in order to determine aquifer hydraulic characteristics representative of a large area. As a rule, 
pumping tests provide more representative measurements of hydraulic parameters; however, they 
require a greater commitment of resources (time, money, and equipment) that cannot be afforded by 
all projects. In addition, for pumping test results to be representative, site hydrogeologic conditions 
should not change appreciably over short distances. This section outlines methods that can be used 
for conducting pump tests in both confined and unconfined aquifers. For a more detailed discussion 
of how to conduct a pumping test, the reader is referred to the work of Dawson and Istok (1991), 
Kruseman and de Ridder (1991), and Driscoll (1986). 

The interpretation of aquifer pumping test data is not unique. Similar sets of data can be ob¬ 
tained from various combinations of geologic conditions. The interpretation of pumping test data is 
discussed in Appendix C of this protocol document. 

A.7.1.1.1 Pumping Test Design 

Prior to performing an aquifer pumping test, all available site and regional hydrogeologic 
information should be assembled and evaluated. Such data should include ground-water flow direc¬ 
tion, hydraulic gradients, other geohydraulic properties, site stratigraphy, well construction details, 
regional water level trends, and the performance of other pumping wells in the vicinity of the test 
area. This information is used to select test duration, proposed pumping rates, and pumping well and 
equipment dimensions. 

The precise location of an aquifer test is chosen to be representative of the area under study. In 
addition, the location is selected on the basis of numerous other criteria, including: 

• Size of the investigation area; 

• Uniformity and homogeneity of the aquifer; 

• Distribution of contaminant sources and dissolved contaminant plumes; 

• Location of known or suspected recharge or barrier boundary conditions; 

• Availability of pumping and/or observation wells of appropriate dimension and screened at 
the desired depth; and 

• Requirements for handling discharge. 

The dimensions and screened interval of the pumping well must be appropriate for the tested 
aquifer. For example, the diameter of the well must be sufficient to accommodate pumping equip- 


A7-26 



ment capable of sustaining the desired flow rate at the given water depth. In addition, if testing a 
confined aquifer that is relatively thin, the pumping well should be screened for the entire thickness 
of the aquifer. For an unconfined aquifer, the wells should be screened in the bottom one-third or 
two-thirds of the saturated zone. 

Any number of observation wells may be used. The number chosen is contingent upon both 
cost and the need to obtain the maximum amount of accurate and reliable data. If three or more 
observation wells are to be installed, and there is a known boundary condition, the wells should be 
placed along a radial line extending from the pumping well toward the boundary, with one well 
placed perpendicular to the line of observation wells to determine whether radial anisotropy exists 
within the aquifer. If two observation wells are to be installed, they should be placed in a triangular 
pattern, non-equidistant from the pumping well. Observation wells should be located at distances 
and depths appropriate for the planned method for analysis of the aquifer test data. Observation well 
spacing should be determined based upon expected drawdown conditions that are the result of the 
studies of geohydraulic properties, proposed pumping test duration, and proposed pumping rate. 
Preliminary pumping results should also be used (if available). Not all projects can afford the luxury 
of preliminary testing. 

The equipment needed to perform aquifer pumping tests includes: 

• Pumps -'Conductivity meter, pH meter, and thermometer 

• Gate valve • Barometer 

• Electrical generator • Semi-log and log-log graph paper 

• Flow meter with totalizer • Portable computer 

• Water level indicators • Field printer for data 

• Pressure gauge • Type matching curves 

• Field logbook/forms • Meter and stopwatch for discharge measurement 

• Pressure transducers and data recorder • Hose or pipe for transfer of water 

• Engineer’s tape calibrated to 0.01 ft • Adequately sized tank for storing contaminated 

• 5-gallon pail water 

Pumping equipment should conform to the size of the well and be capable of delivering the 
estimated range of pumping rates. The selection of flow meter, gate valve, and water transfer lines 
should be based on anticipated rates of water discharge. Both the discharge rate and test duration 
should be considered when selecting a tank for storing discharge water if the water cannot be re¬ 
leased directly to the ground, sanitary sewer, storm sewer, or nearby water treatment facility. 

In areas of severe winter climates, where the frost line may extend to depths of several feet, 
pumping tests should be avoided during cold weather months where the water table is less than 12 
feet from the surface. Under certain conditions, the frozen soil acts as a confining stratum, and 
combined with leaky aquifer and delayed storage characteristics, test results may be unreliable. 

A.7.1.1.2 Preparation for Testing 

Barometric changes may affect water levels in wells, particularly in semiconfined and confined 
aquifers. A change in barometric pressure may cause a change in the water level. Therefore, for at 
least 24 hours prior to performing a pumping test, barometric pressure and water levels in the test 
well, observation wells, and a well beyond the influence of the pumping well should be measured 
hourly to establish trends in ground-water level fluctuation. If a trend is apparent, the barometric 
pressure should be used to develop curves depicting the change in water level versus time. These 
curves should be used to correct the water levels observed during the pumping test. Ground-water 
levels in the background well as well as barometric pressures should continue to be recorded 
throughout the duration of the test. 


A7-27 



Test wells should undergo preliminary pumping or step drawdown tests prior to the actual test. 
This will enable fines to be flushed from the adjacent formation near the well and a steady flow rate 
to be established. The preliminary pumping should determine the maximum drawdown in the well 
and the proper pumping rate should be determined by step drawdown testing. The aquifer should 
then be given time to recover before the actual pumping test begins (as a rule-of-thumb, one day). 

A record should be maintained in the field logbook of the times of pumping and discharge of 
other wells in the area, and if their radii of influence intersect the cone of depression of the test well. 
All measurements and observations should be recorded in a field notebook or on an Aquifer Test 
Data Form. If data loggers with transducers are used, field measurements should be performed in 
case of data logger malfunction. 

A.7.1.1.3 Conducting the Pumping Test 

Immediately prior to starting the pump, the water levels should be measured and recorded for all 
wells to determine the static water levels upon which all drawdowns will be based. Data loggers 
should be reset for each well to a starting water level of 0.0 foot. 

Water pumped from an unconfined aquifer during a pumping test should be disposed of in such 
a manner as not to allow the aquifer to be recharged by infiltration during the test. This means that 
the water must be piped away from the well and associated observation wells. Recharge could 
adversely affect the results. Also, if contaminated water is pumped during the test, the water must be 
stored and treated or disposed of according to the project work plan for the study. The discharge 
water may be temporarily stored in drums, a lined, bermed area, or tanks. If necessary, it should be 
transported and staged in a designated secure area. 

The discharge rate should be measured frequently throughout the test and controlled to maintain 
it as constant as possible, after the initial excess discharge has been stabilized. This can be achieved 
by using a control valve. 

The pitch or rhythm of the pump or generators provides a check on performance. If there is a 
sudden change in pitch, the discharge should be checked immediately and proper adjustments to the 
control valve or the engine speed should be made, if necessary. Do not allow the pump to break 
suction during the test. Allow for maximum drawdown of the well during the step drawdown test. If 
done properly, the flow control valve can be pre-set for the test and will not have to be adjusted 
during pumping. If the pump does shut down during the test, make necessary adjustments and restart 
the test after the well has stabilized. For a confined aquifer, the water level in the pumping well 
should not be allowed, if possible, to fall below the bottom of the upper confining stratum during a 
pumping test. 

At least 10 measurements of drawdown for each log cycle of time should be made both in the 
test well and the observation wells. Data loggers can be set to record in log time, which is very 
useful for data analysis. A suggested schedule for recording water level measurements made by 
hand is as follows: 

• 0 to 10 minutes - 0.5, 1.0, 2.5, 2.0, 2.5, 3.0, 4.5, 6.5, 8, and 10 minutes. It is important in 
the early part of the test to record with maximum accuracy the time at which readings are 
taken. 

• 10 to 100 minutes - 10, 15, 20, 25, 30, 40, 50, 65, 80, and 100 minutes. 

• Then, at 1-hour intervals from 120 minutes to 1,440 minutes (one day) and every 2 hours 
after 1 complete day. 

Initially, there should be sufficient field personnel to station one person at each well used in the 
pumping test (unless an automatic water-level recording system has been installed). After the first 
two hours of pumping, two people are usually sufficient to complete the test. A third person may be 
needed when treatment of the pumped water is required prior to discharge. It is advisable for at least 


A7-28 


one field member to have experience in the performance of pump tests, and for all field personnel to 
have a basic familiarity with conducting the test and gathering data. 

Field personnel should be aware that electronic equipment sometimes fails in the field. Some 
field crews have experienced complete loss of data due to failure of a logger or transducer. It is a 
good idea to record data in the field logbook or on a manual form as the data are produced. That 
way, the data are not lost should the equipment fail. 

The discharge or pumping rate should be measured with a flow meter that also has a totalizer. 
When the pumping is complete, the total gallons pumped are divided by the time of pumping to 
obtain the average discharge rate for the test. Periodic checking and recording of the pumping rate 
during the test also should be performed. 

The total pumping time for a test depends on the type of aquifer and degree of accuracy desired. 
Economizing on the duration of pumping is not recommended. More reliable results are obtained if 
pumping continues until the cone of depression achieves a stabilized condition. The cone of depres¬ 
sion will continue to expand at an ever-decreasing rate until recharge of the aquifer equals the pump¬ 
ing rate, and a steady-state condition is established. The time required for steady-state flow to occur 
may vary from a few hours to years. 

Under normal conditions, it is a good practice to continue a pumping test in a confined aquifer 
for at least 24 hours, and in an unconfined aquifer for a minimum of 72 hours. A longer duration of 
pumping may reveal the presence of boundary conditions or delayed yield. Use of portable comput¬ 
ers allows time/drawdown plots to be made in the field. If data loggers are used to monitor water 
levels, hard copies of the data printed on field printers should be obtained before transporting the 
logger back to the office for downloading. 

A.7.1.2 Slug Tests 

A slug test is a single-well hydraulic test used to determine the hydraulic conductivity of an 
aquifer in the immediate vicinity of the well. Because hydraulic conductivity varies spatially within 
and between aquifers and because slug test results reflect aquifer conditions only in the immediate 
vicinity of the tested well, slug tests should be conducted in as many wells as possible at a site. Slug 
tests can be used for both confined and unconfined aquifers that have transmissivities of less than 
approximately 7,000 square feet per day (ft 2 /day). Slug tests are accomplished by removing a solid 
slug (rising head) or introducing a solid slug (falling head), and then allowing the water level to 
stabilize while taking water level measurements at closely spaced time intervals. The method pre¬ 
sented herein discusses the use of falling head and rising head slug tests in sequence. The analysis of 
slug test data is discussed in Appendix C. 

Slug testing should not proceed until water level measurements show that static water level 
equilibrium has been achieved. Unvented wells should be uncapped at least 24 hours prior to initiat¬ 
ing the test in order to allow the static water level to come to equilibrium. The protective casing 
should remain locked during this time to prevent vandalism. During the slug test, the water level 
change should be influenced only by the introduction or removal of the slug volume. Other factors, 
such as inadequate well development or extended pumping, may lead to inaccurate results. It is the 
field scientist’s responsibility to decide when static equilibrium has been reached in the well. 

The following equipment is needed to conduct a slug test: 

• Teflon®, PVC, or metal slug 

• Nylon or polypropylene rope 

• Electric water level indicator 

• Pressure transducer/sensor 

• Field logbook/forms 

• Automatic data recorder (such as the Hermit Environmental Data Logger®, In-Situ, Inc. 

Model SE1000B, or equal) 


A7-29 



The falling head test is the first step in the two-step slug-testing procedure. The following steps 
describe the recommended falling head slug test procedure: 

1. Decontaminate all downhole equipment. 

2. Record pre-test information including: well number, personnel, climatic data, ground surface 
elevation, measuring point elevation, equipment identifications, and date. 

3. Measure and record the static water level in the well to the nearest 0.01 foot. 

4. Lower the decontaminated pressure transducer into the well and allow the displaced water to 
return to within 0.01 foot of the original static level. 

5. Lower the decontaminated slug into the well to just above the water surface in the well. 

6. Start the data logger and quickly lower the slug below the water table being careful not to 
disturb the pressure transducer. Follow the owner’s manual for proper operation of the data 
logger. 

7. Terminate data recording when the water level has recovered at least 80 percent from the 
initial slug displacement. 

Immediately following completion of the falling head test, the rising head test is performed. 
The following steps describe the rising head slug test procedure: 

1. Measure the static water level in the well to the nearest 0.01 foot to ensure that it has returned 
to the static water level. 

2. Initiate data recording and quickly withdraw the slug from the well. Follow the owner’s 
manual for proper operation of the data logger. 

3. Terminate data recording when the water level has recovered at least 80 percent from the 
initial slug displacement. 

It is advisable to produce hard copies or backup electronic copies of the data logger output (draw¬ 
down vs. time) daily and before transporting the logger from the field site. 

A.7.1.3 Downhole Flow Meter Measurements 

Downhole flow meter measurements are used to investigate the relative vertical distribution of 
horizontal hydraulic conductivity in an open borehole or the screened portion of a well. These 
measurements are useful for identifying zones of elevated hydraulic conductivity that may contribute 
to preferential flow pathways and affect contaminant migration. Methodologies for interpreting data 
from borehole surveys are described by Molz et al. (1994). 

Flowmeter measurements should be perfo/med at 1- to 3-foot intervals in test wells during both» 
ambient conditions and induced flow conditions. Test data may be analyzed using the methods 
described by Molz et al. (1994) to define the relative distribution of horizontal hydraulic conductiv¬ 
ity within the screened interval of each well. Final results should be presented in tabular and graphi¬ 
cal forms and accompanied by appropriate interpretation and discussion. Estimates of bulk hydraulic 
conductivity from previous aquifer tests or results of single-well tests conducted in conjunction with 
the flow meter survey can be used to estimate the absolute hydraulic conductivity distribution at each 
well. 

Borehole flowmeters should be calibrated prior to testing. Generally, 0.5-inch-ID and 1.0-inch- 
ID probes will be calibrated using a range of volumetric flowrates potentially applicable to most sites 
[e.g., approximately 0.04 liters per minute (L/min) to 10 L/min]. The following nine steps outline 
general procedures that can be used to conduct a downhole flow meter survey at a given location. 

• Measure the water level, organic liquid (NAPL) interfaces (if present), and total depth 
(TD) prior to initiating the test. 

• Calibrate the flow meter for the range of anticipated flow velocities before introducing the 
flow meter into the well or borehole. 

• Lower the flow meter to the bottom of the well/borehole. 


A7-30 


• Slowly withdraw the flow meter, pausing to obtain measurements at intervals of approxi¬ 
mately 1 to 3 feet, depending on site conditions. This will provide a baseline under static 
(ambient) conditions. 

• Conduct a short-term, single-well pumping test in the test well to stress the aquifer. 

• Record drawdown using an electronic data logger with a pressure transducer. 

• Monitor and adjust the ground-water extraction rate, as necessary, to maintain constant 
flow. 

• Obtain the profile of the vertical flow at the same elevations occupied during the ambient 
profile upon stabilization of the flow rate. 

• Analyze the data collected during the tests to estimate relative distribution of flow into the 
tested wells and the relative hydraulic conductivity distribution at each location (Molz et 
aU 1994). 

A.7.2 HYDRAULIC GRADIENT 

Hydraulic gradient, defined as the change in ground-water elevation with distance, is a key 
parameter governing the direction and rate of ground-water flow and contaminant migration. Be¬ 
cause ground water can flow in both the horizontal and vertical planes, both horizontal and vertical 
gradients are required for a successful demonstration of monitored natural attenuation. Hydraulic 
gradients are generally calculated on the basis of ground-water elevations measured in site monitor¬ 
ing wells or monitoring points using an electric water level indicator. Therefore, for the most com¬ 
plete representation of site hydrogeology, it is important to measure ground-water elevations from as 
many depths and locations as available. Interpretation of ground-water elevations and the subse¬ 
quent calculations for hydraulic gradient are discussed in Appendix C. 

A.7.3 DIRECT MEASUREMENT OF GROUND-WATER VELOCITY 

Ground-water velocity is directly related to contaminant velocity; therefore, a determination of 
groundwater velocity is critical to the fate and transport portion of a demonstration of monitored 
natural attenuation. Typically, ground-water velocity is estimated from the hydraulic conductivity, 
hydraulic gradient, and effective porosity as described in Appendix C; however, direct measurement 
of ground-water velocity can be obtained from dye tracer studies. 


A7-31 














- 


* 



















APPENDIX B 


IMPORTANT PROCESSES AFFECTING THE FATE AND TRANSPORT OF 

ORGANIC COMPOUNDS IN THE SUBSURFACE 


TABLE OF CONTENTS - APPENDIX B 


B-1 INTRODUCTION.B1 -6 

B.1.1 FATE AND TRANSPORT MECHANISMS.Bl-6 

B. 1.2 MATHEMATICAL DESCRIPTION OF SOLUTE FATE 

AND TRANSPORT.Bl-6 

B-2 NONDESTRUCTIVE ATTENUATION MECHANISMS.B2-9 

B.2.1 ADVECTION.B2-9 

B.2.2 HYDRODYNAMIC DISPERSION.B2-9 

B.2.2.1 Mechanical Dispersion.B2-11 

B.2.2.2 Molecular Diffusion.B2-13 

B.2.2.3 Equation of Hydrodynamic Dispersion.B2-13 

B.2.2.4 One-Dimensional Advection-Dispersion Equation.B2-14 

B.2.3 SORPTION.B2-15 

B.2.3.1 Mechanisms of Sorption.B2-16 

B.2.3.2 Sorption Models and Isotherms.B2-17 

B.2.3.2.1 Langmuir Sorption Model.B2-17 

B.2.3.2.2 Freundlich Sorption Model.B2-18 

B.2.3.3 Distribution Coefficient.B2-19 

B.2.3.4 Coefficient of Retardation.B2-20 

B.2.3.4.1 Determining the Coefficient of Retardation using K oc .B2-20 

B.2.3.4.2 Determining the Coefficient of Retardation using 

Laboratory Tests.B2-24 

B.2.3.5 One-Dimensional Advection-Dispersion Equation with Retardation.B2-25 

B.2.4 VOLATILIZATION.B2-26 

B.2.5 RECHARGE.B2-26 

B-3 DESTRUCTIVE ATTENUATION MECHANISMS - BIOLOGICAL.B3-29 

B.3.1 OVERVIEW OF BIODEGRADATION.B3-30, 

B.3.2 BIODEGRADATION OF ORGANIC COMPOUNDS VIA USE 

AS A PRIMARY GROWTH SUBSTRATE.B3-33 

B.3.2.1 Aerobic Biodegradation of Primary Substrates.B3-33 

B.3.2.1.1 Aerobic Oxidation of Petroleum Hydrocarbons..•.B3-35 

B.3.2.1.2 Aerobic Oxidation of Chlorinated Ethenes.B3-35 

B.3.2.1.3 Aerobic Oxidation of Chlorinated Ethanes.B3-35 

B.3.2.1.4 Aerobic Oxidation of Chlorobenzenes.B3-36 

B.3.2.2 Anaerobic Biodegradation of Primary Substrates.B3-36 

B.3.2.2.1 Anaerobic Oxidation of Petroleum Hydrocarbons.B3-36 

B.3.2.2.2 Anaerobic Oxidation of Chlorinated Ethenes.B3-36 

B.3.2.2.3 Anaerobic Oxidation of Chlorinated Ethanes.B3-36 

B.3.2.2.4 Anaerobic Oxidation of Chlorobenzenes.B3-37 

B.3.3 BIODEGRADATION OF ORGANIC COMPOUNDS VIA USE AS AN 

ELECTRON ACCEPTOR (REDUCTIVE DECHLORINATION).B3-37 

B.3.3.1 Reductive Dechlorination of Chlorinated Ethenes.B3-38 


Bl-2 







































B.3.3.2 Reductive Dechlorination of Chlorinated Ethanes.B3-40 

B.3.3.3 Reductive Dechlorination of Chlorobenzenes.B3-40 

B.3.4 BIODEGRADATION OF ORGANIC COMPOUNDS VIA 

COMETABOLISM.B3-40 

B.3.5 THERMODYNAMIC CONSIDERATIONS.B3-41 

B.3.6 ONE-DIMENSIONAL ADVECTION-DISPERSION EQUATION WITH 

RETARDATION AND BIODEGRADATION.B3-59 

-4 DESTRUCTIVE ATTENUATION MECHANISMS - ABIOTIC.B4-60 

B.4.1 HYDROLYSIS AND DEHYDROHALOGENATION.B4-60 

B.4.1.1 Hydrolysis.B4-60 

B.4.1.2 Dehydrohalogenation.B4-61 

B.4.2 REDUCTION REACTIONS.B4-63 












FIGURES 


No. Title Page 

B.2.1 Breakthrough curve in one dimension showing plug flow with 

continuous source resulting from advection only.B2-10 

B.2.2 Breakthrough curve in one dimension showing plug flow with 

instantaneous source resulting from advection only.B2-10 

B.2.3 Plume migration in two dimensions (plan view) showing plume 
migration resulting from advective flow only with continuous and 

instantaneous source.B2-10 

B.2.4 Physical processes causing mechanical dispersion at the microscopic scale.B2-11 

B.2.5 Breakthrough curve in one dimension showing plug flow with 

instantaneous source resulting from advection only and the combined 

processes of advection and hydrodynamic dispersion.B2-12 

B.2.6 Breakthrough curve in one dimension showing plug flow with 

instantaneous source resulting from advection only and the combined 

processes of advection and hydrodynamic dispersion.B2-12 

B.2.7 Relationship between dispersivity and scale.B2-15 

B.2.8 Breakthrough curve in one dimension showing plug flow with 
continuous source resulting from advection only; the combined 
processes of advection and hydrodynamic dispersion; and the combined 

processes of advection, hydrodynamic dispersion, and sorption.B2-16 

B.2.9 Breakthrough curve in one dimension showing plug flow with 

instantaneous source resulting from advection only; the combined 
processes of advection and hydrodynamic dispersion; and the combined 

processes of advection, hydrodynamic dispersion, and sorption.B2-16 

B.2.10 Characteristic adsorption isotherm shapes.B2-18 

B.2.11 Plot of sorbed concentration versus equilibrium concentration.B2-25 

B.3.1 Breakthrough curve in one dimension showing plug flow with 
continuous source resulting from advection only; the combined 
processes of advection and hydrodynamic dispersion; the combined 
processes of advection, hydrodynamic dispersion, and sorption; and 
the combined processes of advection, hydrodynamic dispersion, 

sorption, and biodegradation...B3-30 

B.3.2 Breakthrough curve in one dimension showing plug flow with 

instantaneous source resulting from advection only; the combined 
processes of advection and hydrodynamic dispersion; the combined 
processes of advection, hydrodynamic dispersion, and sorption; and 
the combined processes of advection, hydrodynamic dispersion, 

sorption, and biodegradation.B3-30 

B.3.3 Oxidation-reduction potentials for various oxidation-reduction reactions.B3-34 

B.3.4 Expected sequence of microbially-mediated redox reactions and 

Gibbs free energy of the reaction.B3-42 


Bl-4 

















TABLES 


No. Title Page 

B. 1.1 Summary of Important Processes Affecting Solute Fate and Transport.B1 -7 

B.2.1 Values of Aqueous Solubility and K oc for Selected Chlorinated Compounds.B2-22 

B.2.2 Values of Aqueous Solubility and for BTEX and Trimethylbenzene Isomers.B2-23 

B.2.3 Data from Hypothetical Batch Test Experiment.B2-25 

B.2.4 Henry’s Law Constants and Vapor Pressures for Common Fuel Hydrocarbons 

and Chlorinated Solvents.B2-27 

B.3.1 Biologic and Abiotic Degradation Mechanisms for Various 

Anthropogenic Organic Compounds.B3-29 

B.3.2 Some Microorganisms Capable of Degrading Organic Compounds.B3-31 

B.3.3 Trends in Contaminant, Electron Acceptor, Metabolic By-product, and Total 

Alkalinity Concentrations During Biodegradation.B3-34 

B.3.4 Sources, Donors, Acceptors, and Products of Reported Reductive 

Dechlorinating Laboratory Systems.B3-39 

B.3.5 Electron Donor and Electron Acceptor Half Cell Reactions.B3-43--B3-44 

B.3.6 Gibbs Free Energy of Formation for Species used in Half Cell Reactions 

and Coupled Oxidation-Reduction Reactions.B3-45--B3-46 

B.3.7 Coupled Oxidation-Reduction Reactions.B3-47--B3-58 

B.4.1 Approximate Half-Lives of Abiotic Hydrolysis and Dehydrohalogenation 

Reactions Involving Chlorinated Solvents.B4-62 


Bl-5 















SECTION B-l 
INTRODUCTION 

B.1.1 FATE AND TRANSPORT MECHANISMS 

This appendix presents an overview of the important processes affecting the fate and transport 
of chlorinated solvents and fuel hydrocarbons dissolved in ground water. The environmental fate 
and transport of a contaminant is controlled by the compound’s physical and chemical properties and 
the nature of the subsurface media through which the compound is migrating. Several processes are 
known to cause a reduction in the concentration and/or mass of a contaminant dissolved in ground 
water. Those processes that result only in the reduction of a contaminant’s concentration but not of 
the total contaminant mass in the system are termed “nondestructive.” Those processes that result in 
degradation of contaminants are referred to as “destructive.” Nondestructive processes include 
advection, hydrodynamic dispersion (mechanical dispersion and diffusion), sorption, dilution, and 
volatilization. Destructive processes include biodegradation and abiotic degradation mechanisms. 
Biodegradation may be the dominant destructive attenuation mechanism acting on a contaminant, 
depending upon the type of contaminant and the availability of electron donors or carbon sources. 
Abiotic degradation processes are also known to degrade chlorinated solvents; where biodegradation 
is not occurring, these may be the only destructive processes operating. However, the rates of abiotic 
processes are generally slow relative to biodegradation rates. 

Remediation by monitored natural attenuation results from the integration of all the subsurface 
attenuation mechanisms (both nondestructive and destructive) operating at a given site. Table B.1.1 
summarizes the processes that affect fate and transport of chlorinated solvents and fuel hydrocarbons 
dissolved in ground water. Important factors to consider include: 

• The compound’s soil/water distribution coefficient (K d ); 

• The compound’s organic carbon/water partition coefficient (K ); 

• The compound’s octanol/water partition coefficient (K ow ); 

• The compound’s water solubility; 

• The compound’s vapor pressure; 

• The compound’s Henry’s Law constant (air/water partition coefficient, H); 

• Indigenous bacterial population; 

• Hydraulic conductivity of aquifer materials; 

• Porosity of aquifer materials; 

• Total organic carbon content of aquifer materials; 

• Bulk density of aquifer materials; 

• Aquifer heterogeneity; and 

• Ambient ground-water geochemistry. 

Nondestructive attenuation mechanisms are discussed in Section B-2. Biodegradation is dis¬ 
cussed in Section B-3. Abiotic degradation mechanisms are discussed in Section B-4. It is impor¬ 
tant to separate nondestructive from destructive attenuation mechanisms during the natural attenua¬ 
tion demonstration. The methods for correcting apparent attenuation caused by nondestructive 
attenuation mechanisms are discussed in Appendix C. 

B.1.2 MATHEMATICAL DESCRIPTION OF SOLUTE FATE AND TRANSPORT 

The partial differential equation describing contaminant migration and attenuation in the 
saturated zone includes terms for advection, dispersion, sorption, and degradation. In one dimen¬ 
sion, the partial differential equation describing solute transport in the saturated zone is: 

dC D x d 2 C v x dC 

dt ~ R dx 2 R dx~^ s eq. B.1.1 


Bl-6 




Table B. 1.1 Summary of Important Processes Affecting Solute Fate and Transport 


Process 

Description 

Dependencies 

Effect 

Advection 

Movement of solute by bulk 
ground-water movement. 

Dependent on aquifer properties, 
mainly hydraulic conductivity and 
effective porosity, and hydraulic 
gradient. Independent of contaminant 
properties. 

Main mechanism driving 
contaminant movement in the 
subsurface. 

Dispersion 

Fluid mixing due to ground- 
water movement and aquifer 
heterogeneities. 

Dependent on aquifer properties and 
scale of observation. Independent of 
contaminant properties. 

Causes longitudinal, transverse, 
and vertical spreading of the 
plume. Reduces solute 
concentration. 

Diffusion 

Spreading and dilution of 
contaminant due to molecular 
diffusion. 

Dependent on contaminant properties 
and concentration gradients. 

Described by Fick’s Laws. 

Diffusion of contaminant from 
areas of relatively high 
concentration to areas of relatively 
low concentration. Generally 
unimportant relative to dispersion 
at most ground-water flow 
velocities. 

Sorption 

Reaction between aquifer matrix 
and solute whereby relatively 
hydrophobic organic compounds 
become sorbed to organic 
carbon or clay minerals. 

Dependent on aquifer matrix 
properties (organic carbon and clay 
mineral content, bulk density, specific 
surface area, and porosity) and 
contaminant properties (solubility, 
hydrophobicity, octanol-water 
partitioning coefficient). 

Tends to reduce apparent solute 
transport velocity and remove 
solutes from the ground water via 
sorption to the aquifer matrix. 

Recharge 
(Simple Dilution) 

Movement of water across the 
water table into the saturated 

zone. 

Dependent on aquifer matrix 
properties, depth to ground water, 
surface water interactions, and 
climate. 

Causes dilution of the contaminant 
plume and may replenish electron 
acceptor concentrations, especially 
dissolved oxygen. 

Volatilization 

Volatilization of contaminants 
dissolved in ground water into 
the vapor phase (soil gas). 

Dependent on the chemical’s vapor 
pressure and Henry’s Law constant. 

Removes contaminants from 
ground water and transfers them to 
soil gas. 

Biodegradation 

Microbially mediated oxidation- 
reduction reactions that degrade 
contaminants. 

Dependent on ground-water 
geochemistry, microbial population 
and contaminant properties. 
Biodegradation can occur under 
aerobic and/or anaerobic conditions. 

May ultimately result in complete 
degradation of contaminants. 
Typically the most important 
process acting to truly reduce 
contaminant mass. 

Abiotic Degradation 

Chemical transformations that 
degrade contaminants without 
microbial facilitation; only 
halogenated compounds are 
subject to these mechanisms in 
the ground-water environment. 

Dependent on contaminant properties 
and ground-water geochemistry. 

Can result in partial or complete 
degradation of contaminants. 

Rates typically much slower than 
for biodegradation. 

Partitioning from 

NAPL 

Partitioning from NAPL into 
ground water. NAPL plumes, 
whether mobile or residual, tend 
to act as a continuing source of 
ground-water contamination. 

Dependent on aquifer matrix and 

contaminant properties, as well as 
ground-water mass flux through or 
past NAPL plume. 

Dissolution of contaminants from 

NAPL represents the primary 
source of dissolved contamination 
in ground water. 


Bl-7 
















Where: 

C = solute concentration [M] 
t = time [T] 

D x = hydrodynamic dispersion [L 2 /T] 

R = coefficient of retardation [dimensionless] 
x = distance along flow path [L] 

= transport velocity in x direction [L/T] 

Q - general source or sink term for reactions involving the 

production or loss of solute (e.g., biodegradation) [M/L 3 /T] 

The degradation of organic contaminants commonly can be approximated using first-order 
kinetics. In one dimension, the partial differential equation describing solute transport with first- 
order decay in the saturated zone is given by: 


dC D x d 2 C v x dC 
dt R dx 2 R dx 


eq. B.1.2 


Where: 

C = concentration [M/L 3 ] 

/ = time [T] 

D x = hydrodynamic dispersion [L 2 /T] 
x = distance along flow path [L] 

R = coefficient of retardation [dimensionless] 

= transport velocity in x direction [L/T] 

X = first-order decay rate [T 1 ] 

These equations serve to illustrate how the processes of advection, dispersion, sorption, and 
biotic and abiotic degradation are integrated to describe the fate and transport of solutes in the 
saturated zone. These relationships were derived using the continuity (conservation of mass) equa¬ 
tion, which states that the rate of change of contaminant mass within a unit volume of porous media 
is equal to the flux of contaminant into the unit volume minus the flux out of the unit volume (Freeze 
and Cherry, 1979). Processes governing flux into the unit volume include advection and hydrody¬ 
namic dispersion (including mechanical dispersion and diffusion). Processes governing flux out of 
the unit volume include advection, hydrodynamic dispersion, dilution, sorption, and chemical reac¬ 
tions (most notably biodegradation). The change in solute concentration may, therefore, be stated 
mathematically as: 

Change in Solute Concentration = Flux In - Flux Out ± Reactions 

The following sections describe the most significant reactions affecting this mass balance (and 
therefore the fate and transport) of organic contaminants in the subsurface. Methods for evaluating 
the flux through the system will be discussed in Appendix C. 


Bl-8 



SECTION B-2 

NONDESTRUCTIVE ATTENUATION MECHANISMS 

B.2.1 ADVECTION 

Advective transport is the transport of solutes by the bulk movement of ground water. Advec- 
tion is the most important process driving dissolved contaminant migration in the subsurface. The 
linear groundwater velocity in the direction parallel to ground-water flow caused by advection is 
given by: 

K dH 

eq. B.2.1 


v * = - 


n. dL 


Where: 

= average linear velocity [L/T] 

K = hydraulic conductivity [L/T] 
n e = effective porosity [L 3 /L 3 ] 
dH/dL - hydraulic gradient [L/L] 

Solute transport by advection alone yields a sharp solute concentration front. Immediately 
ahead of the front, the solute concentration is equal to the background concentration (generally zero). 
At and behind the advancing solute front, the concentration is equal to the initial contaminant con¬ 
centration at the point of release. This is referred to as plug flow and is illustrated in Figures B.2.1, 
B.2.2, and B.2.3. In reality, the advancing front spreads out due to the processes of dispersion and 
diffusion, as discussed in Section B-3, and is retarded by sorption and biodegradation, as discussed 
in Sections B-4 and B-5, respectively. 

The one-dimensional advective transport component of the advection-dispersion equation is 
given by: 

dC dC 

dt ~ V ' dx eq ' B ' 2 ' 2 

Where: 

= average linear velocity [L/T] 

C = contaminant concentration [M/L 3 ] 
t = time [T] 

jc = distance along flow path [L] 

Equation B.2.2 considers only advective transport of the solute. In some cases this may be a 
fair approximation for simulating solute migration because advective transport is the main force 
behind contaminant migration. However, because of dispersion, diffusion, sorption, and biodegrada¬ 
tion, this equation generally must be combined with the other components of the modified advection- 
dispersion equation (equation B. 1.1) to obtain an accurate mathematical description of solute trans¬ 
port. 

B.2.2 HYDRODYNAMIC DISPERSION 

Hydrodynamic dispersion is the process whereby a contaminant plume spreads out in directions 
that are longitudinal and transverse to the direction of plume migration. Dispersion of organic 
solutes in an aquifer is an important consideration when modeling remediation by natural attenua¬ 
tion. Dispersion of a contaminant dilutes the concentrations of the contaminant, and introduces the 
contaminant into relatively pristine portions of the aquifer where it may admix with more electron 
acceptors crossgradient to the direction of ground-water flow. Two very different processes cause 


B2-9 




Contaminant front with 

1.0 

c 
o 

03 

£ Is C/C 0.5 
<u 

■a o 
J2 a 
<u o 
oi u 

0.0 


Figure B.2.1 Breakthrough curve in one dimension showing plug flow with continuous source resulting 
from advection only. 


advection only 



Time or Distance from Source 


a 

_c 


£ 


a 
u 
u 
G 
(U o 

ec U 


CS 


Contaminant front with 



Figure B.2.2 Breakthrough curve in one dimension showing plug flow with instantaneous source resulting 
from advection only. 


Ground-water Flow Direction-► 



Continuous Source 


Instantaneous Source 


Figure B.2.3 Plume migration in two dimensions (plan view) showing plume migration resulting from 
advective flow only with continuous and instantaneous sources. 


B2-10 


















hydrodynamic dispersion; mechanical dispersion and molecular diffusion. The variable describing 
hydrodynamic dispersion, D, is the sum of mechanical dispersion and molecular diffusion. Mechani¬ 
cal dispersion is the dominant mechanism causing hydrodynamic dispersion at normal ground-water 
velocities. At extremely low ground-water velocities, molecular diffusion can become the dominant 
mechanism of hydrodynamic dispersion. Molecular diffusion is generally ignored for most ground- 
water studies. The following sections describe these processes and how they are incorporated into 
the modified advection-dispersion equation (Equation B.1.1). 

B.2.2.1 Mechanical Dispersion 

As defined by Domenico and Schwartz (1990), mechanical dispersion is mixing that occurs as a 
result of local variations in velocity around some mean velocity of flow. With time, a given volume 
of solute will gradually become more dispersed as different portions of the mass are transported at 
the differing velocities. In general, the main cause of variations of both rate and direction of trans¬ 
port velocities is the heterogeneity of the porous aquifer medium. These heterogeneities are present 
at scales ranging from microscopic (e.g., pore to pore) to macroscopic (e.g., well to well) to megas¬ 
copic (e.g., a regional aquifer system). 

Three processes are responsible for mechanical dispersion on the microscopic scale 
(Figure B.2.4). The first process is the variation in flow velocity through pores of various sizes. As 
ground water flows through a porous medium, it flows more slowly through large pores than through 
smaller pores. The second cause of mechanical dispersion is tortuosity, or flow path length. As 
ground water flows through a porous medium, some of the ground water follows less tortuous 
(shorter) paths, while some of the ground water takes more tortuous (longer) paths. The longer the 
flow path, the slower the average linear velocity of the ground water and the dissolved contaminant. 
The final process causing mechanical dispersion is variable friction within an individual pore. 
Groundwater traveling close to the center of a pore experiences less friction than ground water 
traveling next to a mineral grain, and therefore moves faster. These processes cause some of the 
contaminated ground water to move faster than the average linear velocity of the ground water and 
some to move slower. This variation in average velocity of the solute causes dispersion of the 
contaminant. 



Pore Size 


Tortuosity 


Friction in 
Pore Throat 


Figure B.2.4 Physical processes causing mechanical dispersion at the microscopic scale. 

Heterogeneity at the macroscopic and megascopic scales also creates variability in ground water 
and solute velocities, therefore producing dispersion on a larger scale. Geologic features that con- 


B2-11 











tribute to dispersion at the macroscopic scale include stratification characteristics such as changing 
unit geometry, discontinuous units, and contrasting lithologies, and permeability characteristics such 
as nonuniform permeability, directional permeability, and trending permeability (Domenico and 
Schwartz, 1990). Even in aquifer material that appears to be homogeneous, relatively small changes 
in the fraction of fine sediment can change hydraulic conductivity characteristics enough to produce 
significant variations in fluid and solute velocities and thus introduce dispersion. Larger geological 
features will introduce dispersion at the megascopic scale. At this scale, structural features such as 
faults, dipping strata, folds, or contacts will create inhomogeneity, as will stratigraphic features such 
as bedding or other depositional structures. 

As a result of dispersion, the solute front travels at a rate that is faster than would be predicted 
based solely on the average linear velocity of the ground water. The overall result of dispersion is 
spreading and mixing of the contaminant plume with uncontaminated ground water. Figures B.2.5 
and B.2.6 illustrate the effects of hydrodynamic dispersion on an advancing solute front. The com¬ 
ponent of hydrodynamic dispersion contributed by mechanical dispersion is given by the relation¬ 
ship: 

Mechanical Dispersion = a x v x eq. B.2.3 

Where: 

v x = average linear groundwater velocity [L/T] 
a x = dispersivity [L] 

Mechanical dispersion has two components, longitudinal dispersion and transverse (both hori¬ 
zontal and vertical) dispersion. Longitudinal dispersion is the spreading of a solute in a direction 
parallel to the direction of ground-water flow. On the microscopic scale, longitudinal dispersion 



Figure B.2.5 Breakthrough curve in one dimension showing plug flow with continuous source resulting from 
advection only and the combined processes of advection and hydrodynamic dispersion. 


c 

o 

a 


<D 

> 

J3 

o 

PC 


c 

u 

y 

c 

o 

U 



Figure B.2.6 Breakthrough curve in one dimension showing plug flow with instantaneous source resulting 
from advection only and the combined processes of advection and hydrodynamic dispersion. 


B2-12 




















occurs because of velocity changes due to variations in pore size, friction in the pore throat, and 
tortuosity. Transverse dispersion is the spreading of a solute in directions perpendicular to the 
direction of ground-water flow. Transverse dispersion on the microscopic scale is caused by the 
tortuosity of the porous medium, which causes flow paths to branch out from the centerline of the 
contaminant plume. 

B.2.2.2 Molecular Diffusion 

Molecular diffusion occurs when concentration gradients cause solutes to migrate from zones of 
higher concentration to zones of lower concentration, even in the absence of ground-water flow. 
Molecular diffusion is only important at low ground-water velocities, and therefore can be ignored in 
areas with high ground-water velocities (Davis et al ., 1993). 

The molecular diffusion of a solute in ground water is described by Fick’s Laws. Fick’s First 
Law applies to the diffusive flux of a dissolved contaminant under steady-state conditions and, for 
the one-dimensional case, is given by: 

F=-D— eq.B.2.4 

dx 

Where: 

F = mass flux of solute per unit area of time [M/T] 

D = diffusion coefficient (L 2 /T) 

C = solute concentration (M/L 3 ) 
dC 

—r~ = concentration gradient (M/L 3 /L) 


For systems where the dissolved contaminant concentrations are changing with time, Fick’s Second 
Law must be applied. The one-dimensional expression of Fick’s Second Law is: 


dC D d 2 C 
dt dx 2 


eq. B.2.5 


Where: 


dC_ 

dt 


= change in concentration with time [M/T] 


The process of diffusion is slower in porous media than in open water because the ions must 
follow more tortuous flow paths (Fetter, 1988). To account for this, an effective diffusion coeffi¬ 
cient, D*, is used. 

The effective diffusion coefficient is expressed quantitatively as (Fetter, 1988): 


D* = wD 


eq. B.2.6 


Where: 

w = empirical coefficient determined by laboratory experiments [dimensionless] 

The value of w generally ranges from 0.01 to 0.5 (Fetter, 1988). 

B.2.2.3 Equation of Hydrodynamic Dispersion 

Hydrodynamic dispersion, D, has two components, mechanical dispersion and molecular 
diffusion. For one-dimensional flow, hydrodynamic dispersion is represented by the following 
equation (Freeze and Cherry, 1979): 

D x = a x v x +D* eq. B.2.7 


Where: 

D x = longitudinal coefficient of hydrodynamic dispersion in the x direction [L 2 /T] 

a x = longitudinal dispersivity [L] 

v x = average linear ground-water velocity [L/T] 

D* = effective molecular diffusion [L 2 /T] 


B2-13 






Dispersivity is a parameter that is characteristic of the porous medium through which the 
contaminant migrates. Dispersivity represents the spreading of a contaminant over a given length of 
flow, and therefore has units of length. It is now commonly accepted (on the basis of empirical 
evidence) that as the scale of the plume or the system being studied increases, the dispersivity will 
also increase. Therefore, dispersivity is scale-dependent, but at a given scale, data compiled by 
Gelhar et al. (1985 and 1992) show that dispersivity may vary over three orders of magnitude. The 
data of Gelhar et al. (1992) are presented on Figure B.2.7 (with permission from Newell et al., 1996). 

Several approaches can be used to estimate longitudinal dispersivity, a x ., on the field scale (i.e., 
macroscopic to megascopic scales). One technique involves conducting a tracer test. Although this 
is potentially the most reliable method, time and monetary constraints can be prohibitive. Another 
method commonly used to estimate dispersivity when implementing a solute transport model is to 
start with a longitudinal dispersivity of 0.1 times the plume length (Lallemand-Barres and 
Peaudecerf, 1978; Pickens and Grisak, 1981; Spitz and Moreno, 1996). This assumes that 
dispersivity varies linearly with scale. However, Xu and Eckstein (1995) evaluated the same data 
presented by Gelhar et al. (1992) and, by using a weighted least-squares method, developed the 
following relationship for estimating dispersivity: 

a x =0n(Log w L p f ‘" 4 eq. B.2.8 

Where: 

a x = longitudinal dispersivity [L] 

L p = plume length [L] 

Both relationships are shown on Figure B.2.7. In either case, the value derived for dispersivity 
will be an estimate at best, given the great variability in dispersivity for a given plume length. How¬ 
ever, for modeling studies, an initial estimate is needed, and these relationships provide good starting 
points for a modeling study. 

In addition to estimating longitudinal dispersivity, it may be necessary to estimate the transverse 
and vertical dispersivities (a T . and a z ., respectively) for a given site. Several empirical relationships 
between longitudinal dispersivity and transverse and vertical dispersivity have been described. 
Commonly, a T is estimated as 0.1 a x . (based on data from Gelhar et al., 1992), or as 0.33a x . (ASTM, 
1995; US EPA, 1986). Vertical dispersivity (a z ) may be estimated as 0.05a x . (ASTM, 1995), or as 
0.025a x . to 0.1 a x . (US EPA, 1986). 

Some solute transport modelers will start with an accepted literature value for the types of 
materials found in the aquifer matrix. After selecting initial dispersivity values, ihe contaminant 
transport model is calibrated by adjusting the dispersivities (along with other transport parameters, as 
necessary) within the range of accepted literature values until the modeled and observed contaminant 
distribution patterns match (Anderson, 1979). This is a two-step process. The first step is to cali¬ 
brate the flow model to the hydraulic conditions present at the site. After the ground-water flow 
model is calibrated to the hydraulics of the system, the contaminant transport model is calibrated by 
trial and error using various values for dispersivity. There is no unique solution because several 
hydraulic parameters, including hydraulic conductivity, effective porosity, and dispersivity, are 
variable within the flow system (Anderson, 1979; Davis etal., 1993), and other transport parameters 
such as retardation and biodegradation may not be well-defined. 


B.2.2.4 One-Dimensional Advection-Dispersion Equation 

The advection-dispersion equation is obtained by adding hydrodynamic dispersion to 
equation B.2.2. In one dimension, the advection-dispersion equation is given by: 


dC n d 2 C dC 
dt dx 2 dx 


eq. B.2.9 


B2-14 





Source: Newell etal., 1996 


Figure B.2 .7 Relationship between dispersmty and scale. 


Where: 

v x = average linear velocity [L/T] 

C = contaminant concentration [M/L 3 ] 

D x = hydrodynamic dispersion [L 2 /T] 
t - time [T] 

x = distance along flow path [L] 

This equation considers both advection and hydrodynamic dispersion. Because of sorption and 
biodegradation, this equation generally must be combined with the other components of the modified 
advection-dispersion equation presented as equation B. 1.1 to obtain an accurate mathematical de¬ 
scription of solute transport. 

B.2.3 SORPTION 

Many organic contaminants, including chlorinated solvents and BTEX, are removed from 
solution by sorption onto the aquifer matrix. Sorption is the process whereby dissolved contami¬ 
nants partition from the ground water and adhere to the particles comprising the aquifer matrix. 
Sorption of dissolved contamination onto the aquifer matrix results in slowing (retardation) of the 
contaminant relative to the average advective ground-water flow velocity and a reduction in dis¬ 
solved BTEX concentrations in ground water. Sorption can also influence the relative importance of 
volatilization and biodegradation (Lyman et al., 1992). Figures B.2.8 and B.2.9 illustrate the effects 
of sorption on an advancing solute front. 

Keep in mind that sorption is a reversible reaction and that at a given solute concentration, some 
portion of the solute is partitioning to the aquifer matrix and some portion is also desorbing and 
reentering solution. As solute concentrations change, the relative amounts of contaminant that are 
sorbing and desorbing will change. For example, as solute concentrations decrease (perhaps due to 


B2-15 












c 

o 
• «-< 

aJ 

B 

c 
o 
o 
c 
u o 
c< U 


1) 

3 


1.0 


Contaminant front with 
advection only 


C/C 0.5- 


0.0 


Contaminant front with advection, 

^ / Contaminant front 

hydrodynamic dispersion--\ 

/ with advection and 

and sorption \ \ 

hydrodynamic 


V dispersion 


Distance from Source, x 


Figure B.2.8 Breakthrough curve in one dimension showing plug flow with continuous source resulting 

from advection only; the combined processes of advection and hydrodynamic dispersion; and 
the combined processes of advection, hydrodynamic dispersion, and sorption. 



c 


.2 


C3 

t> 

b 

> 

c 

Q 


o 

e 

"3 

o 

OC 

U 


Initial 

contaminant 

slug 

1.0- 


C/C„ 0.5- 



Contaminant slug 
with advection, 
hydrodynamic dispersion, 
and sorption 



Contaminant slug with 
advection only 


Contaminant slug 
with advection 
and hydrodynamic 
dispersion 


Time or Distance from Source 


Figure B.2.9 Breakthrough curve in one dimension showing plug flow with instantaneous source resulting 
from advection only; the combined processes of advection and hydrodynamic dispersion; and 
the combined processes of advection, hydrodynamic dispersion, and sorption. 


plume migration or solute biodegradation and dilution), the amount of contaminant reentering solu- • 
tion will likely increase. The affinity of a given compound for the aquifer matrix will not be suffi¬ 
cient to permanently isolate it from ground water, although for some compounds, the rates of desorp¬ 
tion may be so slow that the loss of mass may be considered permanent for the time scale of interest. 
Sorption, therefore, does not permanently remove solute mass from ground, water; it merely retards 
migration. It is this slowing of contaminant migration that must be understood in order to effectively 
predict the fate of a dissolved contaminant. This section provides information on how retardation 
coefficients are determined in the laboratory. It is not the intent of this document to instruct people 
in how to perform these experiments; this information is provided for informational purposes only. 
Linear isotherms and previously determined soil sorption coefficients (K oc ) are generally used to 
estimate sorption and retardation. 

B.2.3.1 Mechanisms of Sorption 

Sorption of dissolved contaminants is a complex phenomenon caused by several mechanisms, 
including London-van der Waals forces, Coulomb forces, hydrogen bonding, ligand exchange, 
chemisorption (covalent bonding between chemical and aquifer matrix), dipole-dipole forces, dipole- 
induced dipole forces, and hydrophobic forces. Because of their nonpolar molecular structure, 
hydrocarbons most commonly exhibit sorption through the process of hydrophobic bonding. When 


B2-16 















the surfaces comprising the aquifer matrix are less polar than the water molecule, as is generally the 
case, there is a strong tendency for the nonpolar contaminant molecules to partition from the ground 
water and sorb to the aquifer matrix. This phenomenon is referred to as hydrophobic bonding and is 
an important factor controlling the fate of many organic pollutants in soils (Devinny et al., 1990). 
Two components of an aquifer have the greatest effect on sorption: organic matter and clay minerals. 
In most aquifers, the organic fraction tends to control the sorption of organic contaminants. 

B.2.3.2 Sorption Models and Isotherms 

Regardless of the sorption mechanism, it is possible to determine the amount of sorption to be 
expected when a given dissolved contaminant interacts with the materials comprising the aquifer 
matrix. Bench-scale experiments are performed by mixing water-contaminant solutions of various 
concentrations with aquifer materials containing various amounts of organic carbon and clay miner¬ 
als. The solutions are then sealed with no headspace and left until equilibrium between the various 
phases is reached. The amount of contaminant left in solution is then measured. 

Both environmental conservative isotherms (ECI) and constant soil to solution isotherms (CSI) 
can be generated. The ECI study uses the same water concentration but changes the soil to water 
ratio. In CSI isotherm studies, the concentration of contaminant in water is varied while the amount 
of water and sediment is constant. In some instances, actual contaminated water from the site is 
added. Typically, the samples are continually rotated and concentrations measured with time to 
document equilibrium. True equilibrium may require hundreds of hours of incubation but 80 to 90 
percent of equilibrium may be achieved in one or two days. 

The results are commonly expressed as a plot of the concentration of chemical sorbed (pg/g) 
versus the concentration remaining in solution (pg/L). The relationship between the concentration of 
chemical sorbed (C a ) and the concentration remaining in solution (C,) at equilibrium is referred to as 
the sorption isotherm because the experiments are performed at constant temperature. 

Sorption isotherms generally exhibit one of three characteristic shapes depending on the 
sorption mechanism. These isotherms are referred to as the Langmuir isotherm, the Freundlich 
isotherm, and the linear isotherm (a special case of the Freundlich isotherm). Each of these sorption 
isotherms, and related equations, are discussed in the following sections. 


B.2.3.2.1 Langmuir Sorption Model 

The Langmuir model describes sorption in solute transport systems wherein the sorbed con¬ 
centration increases linearly with increasing solute concentration at low concentrations and ap¬ 
proaches a constant value at high concentrations. The sorbed concentration approaches a constant 
value because there are a limited number of sites on the aquifer matrix available for contaminant 
sorption. This relationship is illustrated in Figure B.2.10. The Langmuir equation is described 
mathematically as (Devinny et al ., 1990): 



KC t b 
1 + KC, 


eq. B.2.10 


Where: 

C = sorbed contaminant concentration (mass contaminant/mass soil) 

a 

K = equilibrium constant for the sorption reaction (pg/g) 

C, = dissolved contaminant concentration (pg/ml) 
b = number of sorption sites (maximum amount of sorbed contaminant) 

The Langmuir model is appropriate for highly specific sorption mechanisms where there are a 
limited number of sorption sites. This model predicts a rapid increase in the amount of sorbed 
contaminant as contaminant concentrations increase in a previously pristine area. As sorption sites 
become filled, the amount of sorbed contaminant reaches a maximum level equal to the number of 
sorption sites, b. 


B2-17 



Linear 



Figure B. 2.10 Characteristic adsorption isotherm shapes. 

B.2.3.2.2 Freundlich Sorption Model 

The Langmuir isotherm model can be modified if the number of sorption sites is large (assumed 
infinite) relative to the number of contaminant molecules. This is generally a valid assumption for 
dilute solutions (e.g., downgradient from a petroleum hydrocarbon spill in the dissolved BTEX 
plume) where the number of unoccupied sorption sites is large relative to contaminant concentra¬ 
tions. The Freundlich model is expressed mathematically as (Devinny et al. , 1990): 

C, = K„ cjf” eq. B.2.11 

Where: 

K. = distribution coefficient 

a 

C a - sorbed contaminant concentration (mass contaminant/mass soil, mg/g) 

C x - dissolved concentration (mass contaminant/volume solution, (mg/ml) 
n - chemical-specific coefficient 

The value of n in this equation is a chemical-specific quantity that is determined experimentally. 
Values of 1/n typically range from 0.7 to 1.1, but may be as low as 0.3 and as high as 1.7 (Lyman et 
al 1992). 

The simplest expression of equilibrium sorption is the linear sorption isotherm, a special form 
of the Freundlich isotherm that occurs when the value of n is 1. The linear isotherm is valid for a 
dissolved species that is present at a concentration less than one half of its solubility (Lyman et al ., 
1992). This is a valid assumption for BTEX compounds partitioning from fuel mixtures into ground 
water. Dissolved BTEX concentrations resulting from this type of partitioning are significantly less 
than the pure compound’s solubility in pure water. The linear sorption isotherm is expressed as (Jury 
et al., 1991): 

C,=K d C l eq. B.2.12 

Where: 

K d - distribution coefficient (slope of the isotherm, ml/g). 

C a = sorbed contaminant concentration (mass contaminant/mass soil, pg/g) 

C, = dissolved contaminant concentration (mass contaminant/volume solution, pg/ml) 

The slope of the linear isotherm is the distribution coefficient, K d . 


B2-18 




B.2.3.3 Distribution Coefficient 

The most commonly used method for expressing the distribution of an organic compound 
between the aquifer matrix and the aqueous phase is the distribution coefficient, K d , which is defined 
as the ratio of the sorbed contaminant concentration to the dissolved contaminant concentration: 


£ eq. B.2.13 

Where: 

K d = distribution coefficient (slope of the sorption isotherm, ml/g) 

C a = sorbed concentration (mass contaminant/mass soil or fag/g) 

C l = dissolved concentration (mass contaminant/volume solution or pg/ml) 

The transport and partitioning of a contaminant is strongly dependent on the chemical’s 
soil/water distribution coefficient and water solubility. The distribution coefficient is a measure of 
the sorption/desorption potential and characterizes the tendency of an organic compound to be 
sorbed to the aquifer matrix. The higher the distribution coefficient, the greater the potential for 
sorption to the aquifer matrix. The distribution coefficient is the slope of the sorption isotherm at the 
contaminant concentration of interest. The greater the amount of sorption, the greater the value of K d . 
For systems described by a linear isotherm, K d is a constant. In general terms, the distribution 
coefficient is controlled by the hydrophobicity of the contaminant and the total surface area of the 
aquifer matrix available for sorption. Thus, the distribution coefficient for a single compound will 
vary with the composition of the aquifer matrix. Because of their extremely high specific surface 
areas (ratio of surface area to volume), the organic carbon and clay mineral fractions of the aquifer 
matrix generally present the majority of sorption sites in an aquifer. 

Based on the research efforts of Ciccioli et al. (1980), Karickhoff et al. (1979), and 
Schwarzenbach and Westall (1981), it appears that the primary adsorptive surface for organic chemi¬ 
cals is the organic fraction of the aquifer matrix. However, there is a “critical level of organic mat¬ 
ter” below which sorption onto mineral surfaces is the dominant sorption mechanism (McCarty et 
al., 1981). The critical level of organic matter, below which sorption appears to be dominated by 
mineral-solute interactions, and above which sorption is dominated by organic carbon-solute interac¬ 
tions, is given by (McCarty et al., 1981): 




200 K 


0.84 

ow 


eq. B.2.14 


Where: 

f oc = critical level of organic matter (mass fraction) 

A = surface area of mineralogical component of the aquifer matrix (m 2 /g) 

K ow = octanol-water partitioning coefficient 

From this relationship, it is apparent that the total organic carbon content of the aquifer matrix 
is less important for solutes with low octanol-water partitioning coefficients (KJ. Also apparent is 
the fact that the critical level of organic matter increases as the surface area of the mineralogic 
fraction of the aquifer matrix increases. The surface area of the mineralogic component of the 
aquifer matrix is most strongly influenced by the amount of clay. For compounds with low K ow 
values in materials with a high clay content, sorption to mineral surfaces could be an important factor 
causing retardation of the chemical. 

Several researchers have found that if the distribution coefficient is normalized relative to the 
aquifer matrix total organic carbon content, much of the variation in observed K d values between 
different soils is eliminated (Dragun, 1988). Distribution coefficients normalized to total organic 
carbon content are expressed as K oc . The following equation gives the expression relating K d to K^: 


B2-19 




eq. B.2.15 


Where: 

K = soil sorption coefficient normalized for total organic carbon content 

K. - distribution coefficient 

a 

f oc = fraction total organic carbon (mg organic carbon/mg soil) 

In areas with high clay concentrations and low total organic carbon concentrations, the clay 
minerals become the dominant sorption sites. Under these conditions, the use of to compute K d 
might result in underestimating the importance of sorption in retardation calculations, a source of 
error that will make retardation calculations based on the total organic carbon content of the aquifer 
matrix more conservative. In fact, aquifers that have a high enough hydraulic conductivity to spread 
hydrocarbon contamination generally have low clay content. In these cases, the contribution of 
sorption to mineral surfaces is generally trivial. 

Earlier investigations reported distribution coefficients normalized to total organic matter 
content (K om ). The relationship between f om and f oc is nearly constant and, assuming that the organic 
matter contains approximately 58 percent carbon (Lyman et al ., 1992): 

K oc = \.12AK om eq.B.2.16 


B.2.3.4 Coefficient of Retardation 

As mentioned earlier, sorption tends to slow the transport velocity of contaminants dissolved in 
ground water. The coefficient of retardation, R, is used to estimate the retarded contaminant veloc¬ 
ity. The coefficient of retardation for linear sorption is determined from the distribution coefficient 
using the relationship: 

R = l + 3^± eq.B.2.17 

n 

Where: 

R = coefficient of retardation [dimensionless] 
p 6 = bulk density of aquifer [M/L 3 ] 

K d - distribution coefficient [L 3 /M] 
n - porosity [L 3 /L 3 ] 

The retarded contaminant transport velocity, v c , is given by: 


K = 



eq. B.2.18 


Where: 

v c = retarded contaminant transport velocity [L/T] 
v x = advective ground-water velocity [L/T] 

R = coefficient of retardation [dimensionless] 

Two methods used to quantify the distribution coefficient and amount of sorption (and thus 
retardation) for a given aquifer/contaminant system are presented below. The first method involves 
estimating the distribution coefficient by using for the contaminants and the fraction of organic 
carbon comprising the aquifer matrix. The second method involves conducting batch or column 
tests to determine the distribution coefficient. Because numerous authors have conducted experi¬ 
ments to determine values for common contaminants, literature values are reliable, and it gener¬ 
ally is not necessary to conduct laboratory tests. 


B.2.3.4.1 Determining the Coefficient of Retardation using 

Batch and column tests have been performed for a wide range of contaminant types and concen¬ 
trations and aquifer conditions. Numerous studies have been performed using the results of these 


B2-20 





tests to determine if relationships exist that are capable of predicting the sorption characteristics of a 
chemical based on easily measured parameters. The results of these studies indicate that the amount 
of sorption is strongly dependent on the amount of organic carbon present in the aquifer matrix and 
the degree of hydrophobicity exhibited by the contaminant (Bailey and White, 1970; Karickhoff et 
al ., 1979; Kenaga and Goring, 1980; Brown and Flagg, 1981; Schwarzenbach and Westall, 1981; 
Hassett et al ., 1983; Chiou et al ., 1983). These researchers observed that the distribution coefficient, 
K d , was proportional to the organic carbon fraction of the aquifer times a proportionality constant. 
This proportionality constant, K^, is defined as given by equation B.2.15. In effect, equation B.2.15 
normalizes the distribution coefficient to the amount of organic carbon in the aquifer matrix. Be¬ 
cause it is normalized to organic carbon, values of are dependent only on the properties of the 
compound (not on the type of soil). Values of K oc have been determined for a wide range of chemi¬ 
cals. Table B.2.1 lists K values for selected chlorinated compounds, and Table B.2.2 lists K 
values for BTEX and trimethylbenzene. 

By knowing the value of for a contaminant and the fraction of organic carbon present in the 
aquifer, the distribution coefficient can be determined by using the relationship: 

K d =K 0C f 0C eq. B.2.19 

When using the method presented in this section to predict sorption of the BTEX compounds, 
total organic carbon concentrations obtained from the most transmissive aquifer zone should be 
averaged and used for predicting sorption. This is because the majority of dissolved contaminant 
transport occurs in the most transmissive portions of the aquifer. In addition, because the most 
transmissive aquifer zones generally have the lowest total organic carbon concentrations, the use of 
this value will give a conservative prediction of contaminant sorption and retardation. 


B2-21 


Table B.2.1 


Values of Aqueous Solubility and K oc for Selected Chlorinated Compounds 


Compound 

Solubility (mg/L) 

*oc 

(L/Kg) 

Tetrachloroethene 

150 a 

263 a 

Tetrachloroethene 


359 b 

Tetrachloroethene 

l,503 c 

209 - 238° 

Trichloroethene 

l,100 a 

107 a 

Trichloroethene 


137 b 

Trichloroethene 

l,100 c 

oo 

o 

o 

1,1-Dichloroethene 

2,250 a 

64.6 a 

1,1-Dichloroethene 


80.2 b 

1,1-Dichloroethene 

2,500 d 

150 d 

cis- 1,2-Dichloroethene 


80.2 b 

cis- 1,2-Dichloroethene 

3,500 c 

49° 

trans- 1,2-Dichloroethene 

6,300 a 

58.9 a 

trans-X ,2-Dichloroethene 


80.2° 

trans-X ,2-Dichloroethene 

6,3 00 e 

36° 

Vinyl Chloride 

l,100 a 

2.45 a 

Vinyl Chloride 

2,763 d 

0.4 - 56 d 

1,1,1 -Trichloroethane 

l,495 c 

183° 

1,1,2-Trichloroethane 

4,420 e 

70 e 

1,1-Dichloroethane 

5,060 d 

40 d 

1,2-Dichloroethane 

8,520 c 

33 to 152° 

Chloroethane 

5,710 e 

33 to 143° 

Hexachlorobenzene 

0.006 f 

— 

1,2-Dichlorobenzene 

156 c 

272 - 1480° 

1,3-Dichlorobenzene 

111 8 

293 to 31,600 8 

1,4-Dichlorobenzene 

74 to 87 d 

273 to 1833 d 

Chlorobenzene 

472 d 

83 to 389 d 

Carbon Tetrachloride 

805 8 

110 8 

Chloroform 

7,950 c 

<34° 

Methylene Chloride 

13,000° 

48° 








a From Knox et al., 1993 
b From Jeng et al., 1992; Temperature = 20°C 
c From Howard, 1990; Temperature = 25°C 
d From Howard, 1989; Temperature = 25°C 
e From Howard, 1989; Temperature = 20°C 
f ATSDR, 1990; Temperature = 20°C 
g From Howard, 1990; Temperature = 2(PC 


B2-22 




































Table B.2.2 


Values of Aqueous Solubility and K oc for BTEX and Trimethylbenzene Isomers 


Compound 

Solubility (mg/L) 

*oc 

(L/Kg) 

Benzene 

1750 a 

87. l a 

Benzene 


83 b 

Benzene 

1780 c 

190 c,<u 

Benzene 

1780 c 

62 w 

Benzene 

1780 11 

~tF~ 

Benzene* 

1780 b 

79 hJ ’* 

Benzene 

1780 cjl 

89 k 

Toluene 

515 a 

151 a 

Toluene 


303 b 

Toluene 

537 c 

380 c,d,t 

Toluene 

537° 

110 ce * 

Toluene* 

537 c 

190^' 

Ethylbenzene 

152 a 

158.5 a 

Ethylbenzene 


519 b 

Ethylbenzene 

167S 

680 c,ii 

Ethylbenzene 

167 c 

200 c,e,t 

Ethylbenzene 

140 b 

501 11 ’ 1 

Ethylbenzene* 

~TW~ 

468 hj 

Ethylbenzene 

167° 

398 k 

o-xylene 

152 a 

128.8 a 

o-xylene 


519 b 

o-xylene* 

152 a 

422^’ 

m-xylene 

158 a 


m-xylene 


519 b 

m-xylene 

162 c 

720 c ’ a ’ 1 

m-xylene 

162 c 

210 c,e,t 

m-xylene* 

162 c 

405.37^' 

p-xylene 

198 a 

204 a 

p- xylene 


519 b 

p-xylene* 

198 a 

357^ 

1,2,3 -trimethylbenzene * 

75 

884 b ’* 

1,2,4-trimethylbenzene 

59* 

884 b 

1,2,4-trimethylbenzene* 

59* 

~rnF~ 

1,3,5 -trimethylbenzene * 

72.60 8 

676^’ 


a From Knox et al., 1993 
b From Jeng et al, 1992; Temperature = 20°C 
c From Lyman et al., 1992; Temperature = 25°C 
d Estimatedfrom K ow 
e Estimated from solubility 

f Estimate from solubility generally considered more reliable 
g From Lyman et al., 1992; Temperature = 20°C 
h From Fetter, 1993 

1 Average of 12 equations used to estimate K oc from K ow or K om 
J Average of 5 equations used to estimate K oc from Solubility 

k Average using equations from Kenaga and Goring (1980), Means et al. (1980), and Hassett et al. (1983) to 
estimate K^from solubility 
1 From Sutton and Calder (1975) 

* Recommended value 


B2-23 







































B.2.3.4.2 Determining the Coefficient of Retardation using Laboratory Tests 

The distribution coefficient may be quantified in the laboratory using batch or column tests. 
Batch tests are easier to perform than column tests. Although more difficult to perform, column tests 
generally produce a more accurate representation of field conditions than batch tests because con¬ 
tinuous flow is involved. Knox et al. (1993) suggest using batch tests as a preliminary screening 
tool, followed by column studies to confirm the results of batch testing. The authors of this docu¬ 
ment feel that batch tests, if conducted properly, will yield sufficiently accurate results for fate and 
transport modeling purposes provided that sensitivity analyses for retardation are conducted during 
the modeling. 

Batch testing involves adding uncontaminated aquifer material to a number of vessels, adding 
solutions prepared using uncontaminated ground water from the site mixed with various amounts of 
contaminants to produce varying solute concentrations, sealing the vessel and shaking it until equi¬ 
librium is reached, analyzing the solute concentration remaining in solution, and calculating the 
amount of contaminant sorbed to the aquifer matrix using mass balance calculations. A plot of the 
concentration of contaminant sorbed versus dissolved equilibrium concentration is then made using 
the data for each reaction vessel. The slope of the line formed by connecting each data point is the 
distribution coefficient. The temperature should be held constant during the batch test, and should 
approximate that of the aquifer system through which solute transport is taking place. 

Table B.2.3 contains data from a hypothetical batch test. These data are plotted (Figure B.2.11) 
to obtain an isotherm unique to the aquifer conditions at the site. A regression analysis can then be 
performed on these data to determine the distribution coefficient. For linear isotherms, the distribu¬ 
tion coefficient is simply the slope of the isotherm. In this example, K d = 0.0146 L/g. Batch-testing 
procedures are described in detail by Roy et al. (1992). 

Column testing involves placing uncontaminated aquifer matrix material in a laboratory column 
and passing solutions through the column. Solutions are prepared by mixing uncontaminated ground 
water from the site with the contaminants of interest and a conservative tracer. Flow rate and time 
are accounted for and samples are periodically taken from the effluent of the column and analyzed to 
determine contaminant and tracer concentrations. Breakthrough curves are prepared for the contami¬ 
nants by plotting chemical concentration versus time (or relative concentration versus number of 
pore volumes). The simplest way to determine the coefficient of retardation (or the distribution 
coefficient) from the breakthrough curves is to determine the time required for the effluent concen¬ 
tration to equal 0.5 of the influent concentration. This value can be used to determine average 
velocity of the center of mass of the contaminant. The retardation factor is determined by dividing 
the average flow velocity through the column by the velocity of the center of mass of the contami¬ 
nant. The value thus obtained is the retardation factor. The coefficient of retardation also can be 
determined by curve fitting using the CXTFIT model of Parker and van Genuchten (1984). Break¬ 
through curves also can be made for the conservative tracer. These curves can be used to determine 
the coefficient of dispersion by curve fitting using the model of Parker and van Genuchten (1984). 

When using the method presented in this section to predict sorption of the BTEX compounds, 
aquifer samples should be obtained from the most transmissive aquifer zone. This is because the 
majority of dissolved contaminant transport occurs in the most transmissive portions of the aquifer. 

In addition, because the most transmissive aquifer zones generally have the lowest organic carbon 
concentrations, the use of these materials will give a conservative prediction of contaminant sorption 
and retardation. 


B2-24 


Table B.2.3 Data from Hypothetical Batch Test Experiment 


Initial Concentration 
(pft/L) 

Equilibrium Concentration 

Weight of Solid 
Matrix (g) 

Sorbed Concentration* (j^g/g) 

250 

77.3 

20.42 

1.69 

500 

150.57 

20.42 

3.42 

1000 

297.04 

20.42 

6.89 

1500 

510.1 

20.42 

9.70 

2000 

603.05 

20.42 

13.68 

3800 

1198.7 

20.42 

25.48 

6000 

2300.5 

20.42 

36.23 

9000 

3560.7 

20.42 

53.27 


* Adsorbed concentration = ((Initial concentration - Equilibrium Concentration) x Volume of Solution) / Weight of 
Solid Matrix 



Figure B.2.11 Plot of sorbed concentration vs. equilibrium concentration. 


B.2.3.5 One-Dimensional Advection-Dispersion Equation with Retardation 

The advection-dispersion equation is obtained by adding hydrodynamic dispersion to 
equation B.2.2. In one dimension, the advection-dispersion equation is given by: 


„sc „ e 2 c ec 

dt dx 2 dx 


eq. B.2.20 


Where: 

v = average linear velocity ground-water velocity [L/T] 

R = coefficient of retardation [dimensionless] 

C = contaminant concentration [M/L 3 ] 

D = hydrodynamic dispersion [L 2 /T] 
t = time [T] 

x = distance along flow path [L] 

This equation considers advection, hydrodynamic dispersion, and sorption (retardation). Be¬ 
cause of biodegradation, this equation generally must be combined with the other components of the 
modified advection-dispersion equation, presented as equation B. 1.1, to obtain an accurate math¬ 
ematical description of solute transport. 


B2-25 




















B.2.4 VOLATILIZATION 

While not a destructive attenuation mechanism, volatilization does remove contaminants from 
the ground-water system. In general, factors affecting the volatilization of contaminants from ground 
water into soil gas include the contaminant concentration, the change in contaminant concentration 
with depth, the Henry’s Law constant and diffusion coefficient of the compound, mass transport 
coefficients for the contaminant in both water and soil gas, sorption, and the temperature of the water 
(Larson and Weber, 1994). 

Partitioning of a contaminant between the liquid phase and the gaseous phase is governed by 
Henry’s Law. Thus, the Henry’s Law constant of a chemical determines the tendency of a contami¬ 
nant to volatilize from ground water into the soil gas. Henry’s Law states that the concentration of a 
contaminant in the gaseous phase is directly proportional to the compound’s concentration in the 
liquid phase and is a constant characteristic of the compound. Stated mathematically, Henry’s Law is 
given by (Lyman et al ., 1992): 

C a = HC, eq. B.2.21 

Where: 

H - Henry’s Law Constant (atm m 3 /mol) 

C a = concentration in air (atm) 

C t = concentration in water (mol/m 3 ) 

Henry’s Law constants for chlorinated and petroleum hydrocarbons range over several orders of 
magnitude. For petroleum hydrocarbons, Henry’s Law constants (H) for the saturated aliphatics, 

H range from 1 to 10 atm m 3 /mol @ 25°C; for the unsaturated and cyclo-aliphatics ranges from 0.1 to 
1 atm m 3 /mol @ 25°C; and for the light aromatics (e.g., BTEX) H ranges from 0.007 to 
0.02 atm m 3 /mol @ 25°C (Lyman et al., 1992). Values of Henry’s Law constants for selected chlori¬ 
nated solvents and the BTEX compounds are given in Table B.2.4. As indicated on the table, values 
of H for chlorinated compounds also vary over several orders of magnitude, although most are 
similar to those for BTEX compounds. 

The physiochemical properties of chlorinated solvents and the BTEX compounds give them low 
Henry’s Law constants, with the exception of vinyl chloride. Because of the small surface area of the 
ground-water flow system exposed to soil gas, volatilization of chlorinated solvents and BTEX 
compounds from ground water is a relatively slow process that, in the interest of being conservative, 
generally can be neglected when modeling biodegradation. Chiang et al. (1989) demonstrated that 
less than 5 percent of the mass of dissolved BTEX is lost to volatilization in the saturated ground- 
water environment. Moreover, Rivett (1995) observed that for plumes more than about 1 meter 
below the air-water interface, little, if any, solvent concentrations will be detectable in soil gas due to 
the downward ground-water velocity in the vicinity of the water table. This suggests that for por¬ 
tions of plumes more than 1 meter below the water table, very little, if any, mass will be lost due to 
volatilization. In addition, vapor transport across the capillary fringe can be very slow (McCarthy 
and Johnson, 1993), thus further limiting mass transfer rates. Because of this, the impact of volatil¬ 
ization on dissolved contaminant reduction can generally be neglected, except possibly in the case of 
vinyl chloride. However, Rivett’s (1995) findings should be kept in mind even when considering 
volatilization as a mechanism for removal of vinyl chloride from ground water. 

B.2.5 RECHARGE 

Groundwater recharge can be defined as the entry into the saturated zone of water made avail¬ 
able at the water-table surface (Freeze and Cherry, 1979). In recharge areas, flow near the water 
table is generally downward. Recharge defined in this manner may therefore include not only pre¬ 
cipitation that infiltrates through the vadose zone, but water entering the ground-water system due to 
discharge from surface water bodies (i.e., streams and lakes). Where a surface water body is in 


B2-26 


Table B.2.4 Henry s Law Constants and Vapor Pressures for Common Fuel Hydrocarbons and 

Chlorinated Solvents 


Compound 

Vapor Pressure (mmHg 
@ 25°C) 

Henry’s Law Constant 

(atm-m^/mol) 

Benzene 

95 

0.0054 

Ethylbenzene 

10 

0.0066 

Toluene 

28.4 

0.0067 

o-Xylene 

10 

0.00527 

m-Xylene 

10 

0.007 

/7-Xylene 

10 

0.0071 

7,2,5-Trimethylbenzene 


0.00318 

7,2,4-Trimethylbenzene 


0.007 

7,5,5-Trimethylbenzene 


0.006 

1,2,4, 5-Tetramethylbenzene 


0.0249 

Tetrachloroethene 

14 

0.0153 

Trichloroethene 

57.8 

0.0091 

1,1-Dichloroethene 

591 

0.018 

m-1,2-Dichloroethene 

200 

0.0037 

trans-\ ,2-Dichloroethene 

265 

0.0072 

Vinyl Chloride 

2,580 

1.22 

1,1,1 -Trichloroethane 

123.7 

0.008 

1,1,2-Trichloroethane 

30.3 

0.0012 

1,1-Dichloroethane 

227 

0.0059 

1,2-Dichloroethane 

78.7 

0.00098 

Chi oroe thane 

766 

0.0085 

Hexachlorobenzene 

0.0000109 

0.00068 

1,2-Dichlorobenzene 

1.47 

0.0012 

1,3-Dichlorobenzene 

2.3 

0.0018 

1,4-Dichlorobenzene 

1.76 

0.0015 

Chlorobenzene 

11.9 

0.0035 

Carbon Tetrachloride 

113.8 

0.0304 

Chloroform 

246 

0.00435 

Methylene Chloride 

434.9 

0.00268 


B2-27 


































contact with or is part of the ground-water system, the definition of recharge above is stretched 
slightly. However, such bodies are often referred to as recharging lakes or streams. Recharge of a 
water table aquifer has two effects on the natural attenuation of a dissolved contaminant plume. 
Additional water entering the system due to infiltration of precipitation or from surface water will 
contribute to dilution of the plume, and the influx of relatively fresh, electron-acceptor-charged water 
will alter geochemical processes and in some cases facilitate additional biodegradation. 

Recharge from infiltrating precipitation is the result of a complex series of processes in the 
unsaturated zone. Description of these processes is beyond the scope of this discussion; however, it 
is worth noting that the infiltration of precipitation through the vadose zone brings the water into 
contact with the soil and thus may allow dissolution of additional electron acceptors and possibly 
organic soil matter (a potential source of electron donors). Infiltration, therefore, provides fluxes of 
water, inorganic species, and possibly organic species into the ground water. Recharge from surface 
water bodies occurs when the hydraulic head of the body is greater than that of the adjacent ground 
water. The surface water may be a connected part of the ground-water system, or it may be perched 
above the water table. In either case, the water entering the ground-water system will not only aid in 
dilution of a contaminant plume but it may also add electron acceptors and possibly electron donors 
to the ground water. 

An influx of electron acceptors will tend to increase the overall electron-accepting capacity 
within the contaminant plume. In addition to the inorganic electron acceptors that may be dissolved 
in the recharge (e.g., dissolved oxygen, nitrate, or sulfate), the introduction of water with different 
geochemical properties may foster geochemical changes in the aquifer. For example, iron (II) will be 
oxidized back to iron (III). Vroblesky and Chapelle (1994) present data from a site where a major 
rainfall event introduced sufficient dissolved oxygen into the contaminated zone to cause 
reprecipitation of iron (M) onto mineral grains. This reprecipitation made iron (HI) available for 
reduction by microorganisms, thus resulting in a shift from methanogenesis back to iron (El) reduc¬ 
tion (Vroblesky and Chapelle, 1994). Such a shift may be beneficial for biodegradation of com¬ 
pounds used as electron donors, such as fuel hydrocarbons or vinyl chloride. However, these shifts 
can also make conditions less favorable for reductive dehalogenation. 

Evaluating the effects of recharge is typically difficult. The effects of dilution might be esti¬ 
mated if one has a detailed water budget for the system in question, but if a plume has a significant 
vertical extent, it cannot be known with any certainty what proportion of the plume mass is being 

z * 

diluted by the recharge. Moreover, because dispersivity, sorption, and biodegradation are often not 
well-quantified, separating out the effects of dilution may be very difficult indeed. Where recharge 
enters from precipitation, the effects of the addition of electron acceptors may be qualitatively appar¬ 
ent due to elevated electron acceptor concentrations or differing patterns in electron acceptor con¬ 
sumption or byproduct formation in the area of the recharge. However, the effects of short-term 
variations in such a system (which are likely due to the intermittent nature of precipitation events in 
most climates) may not be easily understood. Where recharge enters from surface water, the influx 
of mass and electron acceptors is more steady over time. Quantifying the effects of dilution may be 
less uncertain, and the effects of electron acceptor replenishment may be more easily identified 
(though not necessarily quantified). 


B2-28 


SECTION B-3 

DESTRUCTIVE ATTENUATION MECHANISMS - BIOLOGICAL 

Many anthropogenic organic compounds, including certain chlorinated solvents, can be de¬ 
graded by both biological and abiotic mechanisms. Biological degradation mechanisms are dis¬ 
cussed in this section; abiotic degradation mechanisms are discussed in Section B.4. Table B.3.1 
summarizes the various biotic and abiotic mechanisms that result in the degradation of anthropo¬ 
genic organic compounds. Biological degradation mechanisms tend to dominate in most ground- 
water systems, depending on the type of contaminant and the ground-water chemistry. 

Table B.3.1 Biologic and Abiotic Degradation Mechanisms for Various Anthropogenic Organic 

Compounds 


Compound 

Degradation Mechanism 

PCE 

Reductive dechlorination 

TCE 

Reductive d'echlorination, cometabolism 

DCE 

Reductive dechlorination, direct biological oxidation 

Vinyl Chloride 

Reductive dechlorination, direct biological oxidation 

TCA 

Reductive dechlorination, hydrolysis, 
dehydrohalogenation 

1,2-DCA 

Reductive dechlorination, direct biological oxidation 

Chloroethane 

Hydrolysis 

Carbon Tetrachloride 

Reductive dechlorination, cometabolism, abiotic 

Chloroform 

Reductive dechlorination, cometabolism 

Methylene Chloride 

Direct biological oxidation 

Chlorobenzenes 

Direct biological oxidation, reductive dechlorination, 
cometabolism 

Benzene 

Direct biological oxidation 

Toluene 

Direct biological oxidation 

Ethylbenzene 

Direct biological oxidation 

Xylenes 

Direct biological oxidation 

1,2-Dibromoethane 

Reductive dehalogenation, hydrolysis, direct 
biological oxidation 


Many organic contaminants are biodegraded by microorganisms indigenous to the subsurface 
environment. During biodegradation, dissolved contaminants are ultimately transformed into in¬ 
nocuous byproducts such as carbon dioxide, chloride, methane, and water. In some cases, intermedi¬ 
ate products of these transformations may be more hazardous than the original compound; however, 
they may also be more easily degraded. Biodegradation of organic compounds dissolved in ground 
water results in a reduction in contaminant concentration (and mass) and slowing of the contaminant 
front relative to the average advective ground-water flow velocity. Figures B.3.1 and B.3.2 illustrate 
the effects of biodegradation on an advancing solute front. 


B3-29 






















c 

o 

1 
5 
o 
o 
a 
o o 
as U 


o 

w 

03 


1 . 0 , 


O.CH 


Contaminant front with advection 
hydrodynamic dispersion 

C/Q-, 0.5- sorption.. 

Contaminant front with advection, 
hydrodynamic dispersion, 
sorption, and biodegradation 



0 


Contaminant front with 
advection only 


Contaminant front 
with advection and 
hydrodynamic 
dispersion 


Distance from Source, x 


Figure B.3.1 Breakthrough curve in one dimension showing plug flow with continuous source resulting 

from advection only; the combined processes of advection and hydrodynamic dispersion; the 
combined processes of advection, hydrodynamic dispersion, and sorption; and the combined 
processes of advection, hydrodynamic dispersion, sorption, and biodegradation. 


a 

o 


u I 
> g 


Initial 

contaminant 

slug 


Contaminant slug 
with advection, 


Contaminant slug with 


C/C n 0.5 


JS a 
o o 
C* O 


O.OJ 



and sorption \ 


auvvvuuii uiiij 


Contaminant slug \ 




with advection, 


y . Contaminant slug 

- 

hydrodynamic dispersion, / 


\ with advection 


sorption, and if _X\ J 


Vf and hydrodynamic 


biodegradation— ^ // V\ J 


\ dispersion 


Time or Distance from Source 


Figure B.3.2 Breakthrough curve in one dimension showing plug flow with instantaneous source resulting 

from advection only; the combined processes of advection and hydrodynamic dispersion; the 
combined processes of advection, hydrodynamic dispersion, and sorption; and the combined 
processes of advection, hydrodynamic dispersion, sorption, and biodegradation. 


B.3.1 OVERVIEW OF BIODEGRADATION 

As recently as 1975 the scientific literature reported the subsurface/aquifer environment as 
devoid of significant biological activity. It is now known that soils and shallow sediments contain a 
large variety of microorganisms, ranging from simple prokaryotic bacteria and cyanobacteria to more 
complex eukaryotic algae, fungi, and protozoa. Over the past two decades, numerous laboratory and 
field studies have shown that microorganisms indigenous to the subsurface environment can degrade 
a variety of organic compounds, including components of gasoline, kerosene, diesel, jet fuel, chlori¬ 
nated ethenes, chlorinated ethanes, the chlorobenzenes, and many other compounds (e.g., for fuels 
see Jamison et ah , 1975; Atlas, 1981, 1984, and 1988; Young, 1984; Bartha, 1986; B. H. Wilson et 
ah , 1986 and 1990; Barker etal., 1987; Baedecker etal., 1988; Lee, 1988; Chiang etal., 1989; 
Cozzarelli et ah , 1990; Leahy and Colewell, 1990; Alvarez and Vogel, 1991; Evans et ah , 1991a and 
1991b; Edwards et ah , 1992; Edwards and Grbic-Galic, 1992; Thierrin et ah , 1992; Malone et ah, 
1993; Davis et ah , 1994a and 1994b; and Lovley et ah , 1995; and for chlorinated solvents see 
Brunner and Leisinger, 1978; Brunner et ah , 1980; Rittman and McCarty, 1980; Bouwer et ah , 1981; 


B3-30 



















Table B.3.2 Some Microorganisms Capable of Degrading Organic Compounds (Modified from Riser- 
Roberts, 1992) 


Contaminant 

Microorganisms 

Comments/ 

Biodegradability 

Benzene 

Pseudomonas putida, P. rhodochrous, P. aeruginosa, 
Acinetobacter sp., Methylosinus trichosporium OB3b, Nocardia 
sp., methanogens, anaerobes 

Moderate to High 

Toluene 

Methylosinus trichosporium OB3b, Bacillus sp., Pseudomonas 
sp., P. putida, Cunninghamella elegans, P. aeruginosa, P. 
mildenberger, P. aeruginosa, Achromobacter sp., methanogens, 
anaerobes 

High 

Ethylbenzene 

Pseudomonas putida 

High 

Xylenes 

Pseudomonas putida, methanogens, anaerobes 

High 

Jet Fuels 

Cladosporium, Hormodendrum 

High 

Kerosene 

Torulopsis, Candidatropicalis, Corynebacterium 
hydrocarboclastus, Candidaparapsilosis, C. guilliermondii, C. 
lipolytica, Trichosporon sp., Rhofiosporidium toruloides, 
Cladosporium resinae 

High 

Chlorinated 

Ethenes 

Dehalobacter restrictus, Dehalospirillum multivorans, 
Enterobacter agglomerans, Dehalococcus entheogenes strain 

195 ,Desulfitobacterium sp. strain PCE1, Pseudomonas putida 
(multiple strains), P. cepacia G4, P. mendocina, 
Desulfobacterium sp., Methanobacterium sp., Methanosarcina 
sp. strain DCM, Alcaligenes eutrophus JMP 134, Methylosinus 
trichosporium OB3b, Escherichia coli, Nitorsomonas europaea, 
Methylocystisparvus OBBP, Mycobacterium sp., Rhodococcus 
erythopolis 

Moderate 

Chlorinated 

Ethanes 

Desulfobacterium sp., Methanobacterium sp., Pseudomonas 
putida, Clostridium sp., C. sp. strain TCAIIB, 

Moderate 

Chlorinated 

Methanes 

Acetobacterium xvoodii, Desulfobacterium sp., 
Methanobacterium sp., Pseudomonas sp. strain KC, Escherichia 
coli K-12, Clostridium sp., Methanosarcina sp., 
Hyphomicrobium sp. strain DM2, 

Moderate 

Chlorobenzenes 

Alcaligenes sp. (multiple strains), Pseudomonas sp. (multiple 
strains,), P. putida, Staphylococcus epidermis 

Moderate to High 


Miller and Guengerich, 1982; Roberts et al., 1982; Bouwer and McCarty, 1983; Stucki et al., 1983; 
Reineke and Knackmuss, 1984; Wilson and Wilson, 1985; Fogel et al., 1986; Egli et al., 1987; Vogel 
and McCarty, 1987; Vogel et al., 1987; Bouwer and Wright, 1988; Little et al., 1988; Freedman and 
Gossett, 1989; Sewell and Gibson, 1991; Chapelle, 1993; DeBruin et al., 1992; Ramanand etal., 
1993; Vogel, 1994; Suflita and Townsend, 1995; Adriaens and Vogel, 1995; Bradley and Chapelle, 
1996; Gossett and Zinder, 1996; Spain, 1996). Table B.3.2 presents a partial list of microorganisms 
known to degrade anthropogenic organic compounds. 

Although we now recognize that microorganisms are ubiquitous in drinking water aquifers, the 
study of the microbial ecology and physiology of the subsurface, below the rhizosphere, is still in its 
infancy. However, great progress has been made at least in identifying, if not fully understanding, 


B3-31 
















the numerous and diverse types of microbially-mediated contaminant transformations that can occur 
in the subsurface. 

Chemothrophic organisms, such as humans and most microorganisms, obtain energy for growth 
and activity from physiologically coupling oxidation and reduction reactions and harvesting the 
chemical energy that is available. Under aerobic conditions (in the presence of molecular oxygen) 
humans and many bacteria couple the oxidation of organic compounds (food) to the reduction of 
oxygen (from the air). However in the absence of oxygen (anaerobic conditions), microorganisms 
may use other compounds as electron acceptors. Anaerobic microorganisms can obtain energy from 
a variety of electron donors such as natural organic carbon or many forms of anthropogenic carbon 
and electron acceptors such as nitrate, iron (ID), sulfate, carbon dioxide, as well as many of the 
chlorinated solvents. 

The introduction of oxidizable soluble organic contaminants into ground water initiates a series 
of complex responses by subsurface microorganisms. Field and laboratory research suggests that 
distinct communities defined by the dominant electron acceptor develop which are spatially and 
temporally separate. These communities are most likely ecologically defined by the flux of biologi¬ 
cally available electron donors and acceptors. The biological processes of these communities are 
potentially useful as natural attenuation mechanisms, as the basis of new bioremediation technolo¬ 
gies, and as indicators of the extent and severity of the release. As electron acceptors and nutrients 
are depleted by microbial activity during biodegradation of contaminants, the redox potential of 
contaminated aquifers decreases. This results in a succession of bacterial types adapted to specific 
redox regimes and electron acceptors. Metabolic byproducts of contaminant biodegradation also 
exert selective forces, either by presenting different carbon sources or by further modifying the 
physical and chemical environment of the aquifer. Like organic and inorganic colloids, microorgan¬ 
isms possess complex surface chemistry, and can themselves serve as mobile and immobile reactive 
sites for contaminants. 

Under anaerobic conditions, most organic compounds are degraded by groups of interacting 
microorganisms referred to as a consortium. In the consortium, individual types of organisms carry 
out different specialized reactions which, when combined, can lead to the complete mineralization of 
a particular compound. The metabolic interaction between organisms can be complex and may be so 
tightly linked under a given set of conditions that stable consortia can be mistakenly identified as a 
single species. There seems to be several advantages to the consortial system, including: 1) This 
system allows for the creation of microenvironments where certain types of organisms can survive in 
otherwise hostile conditions; 2) Reactions that are thermodynamically unfavorable can be driven by 
favorable reactions when they are metabolically linked within the consortium; and, 3) This system 
takes advantage of the diverse metabolic capabilities of microorganisms by allowing for the forma¬ 
tion and enrichment of associations that can utilize an introduced substrate faster than a single 
species could evolve a novel complex enzyme pathway to degrade the same compound. 

It appears that subsurface microbial communities contain the metabolic diversity required to 
utilize a wide variety of organic contaminants as a primary growth substrate in the presence of 
electron acceptors such as oxygen. Some pollutants, especially the highly oxidized chlorinated 
hydrocarbons, are not amenable to use as a primary growth substrate. Instead, these compounds are 
used as electron acceptors in reactions that rely on another source of carbon as a primary substrate or 
are degraded fortuitously via cometabolism. Thus, biodegradation of organic compounds in ground 
water occurs via three mechanisms: 

• Use of the organic compound as the primary growth substrate; 

• Use of the organic compound as an electron acceptor; and 

• Cometabolism. 


B3-32 


The first two biodegradation mechanisms involve the microbial transfer of electrons from 
electron donors (primary growth substrate) to electron acceptors. This process can occur under 
aerobic or anaerobic conditions. Electron donors include natural organic material, fuel hydrocar¬ 
bons, chlorobenzenes, and the less oxidized chlorinated ethenes and ethanes. Electron acceptors are 
elements or compounds that occur in relatively oxidized states. The most common naturally occur¬ 
ring electron acceptors in ground water include dissolved oxygen, nitrate, manganese (IV), iron (III), 
sulfate, and carbon dioxide. In addition, the more oxidized chlorinated solvents such as PCE, TCE, 
DCE, TCA, DCA, and polychlorinated benzenes can act as electron acceptors under favorable 
conditions. Under aerobic conditions, dissolved oxygen is used as the terminal electron acceptor 
during aerobic respiration. Under anaerobic conditions, the electron acceptors listed above are used 
during denitrification, manganese (IV) reduction, iron (III) reduction, sulfate reduction, 
methanogenesis, or reductive dechlorination. Chapelle (1993) and Atlas (1988) discuss terminal 
electron accepting processes in detail. 

The third biodegradation mechanism is cometabolism. During cometabolism the compound 
being degraded does not benefit the organism. Instead, degradation is brought about by a fortuitous 
reaction wherein an enzyme produced during an unrelated reaction degrades the organic compound. 

As discussed in sections B.3.2, B.3.3, and B.3.4, biodegradation causes measurable changes in 
ground-water chemistry. Table B.3.3 summarizesthese trends. During aerobic respiration, oxygen is 
reduced to water, and dissolved oxygen concentrations decrease. In anaerobic systems where nitrate 
is the electron acceptor, the nitrate is reduced to N0 2 , N 2 0, NO, NH 4+ , or N 2 via denitrification or 
dissimilatoiy nitrate reduction, nitrate concentrations decrease. In anaerobic systems where iron (III) 
is the electron acceptor, it is reduced to iron (II) via iron (IE) reduction, and iron (II) concentrations 
increase. In anaerobic systems where sulfate is the electron acceptor, it is reduced to H 2 S via sulfate 
reduction, and sulfate concentrations decrease. During aerobic respiration, denitrification, iron (III) 
reduction, and sulfate reduction, total alkalinity will increase. In anaerobic systems where C0 2 is 
used as an electron acceptor, it is reduced by methanogenic bacteria during methanogenesis, and CH 4 
is produced. In anaerobic systems where contaminants are being used as electron acceptors, they are 
reduced to less chlorinated daughter products; in such a system, parent compound concentrations 
will decrease and daughter product concentrations will increase at first and then decrease as the 
daughter product is used as an electron acceptor or is oxidized. 

As each subsequent electron acceptor is utilized, the ground water becomes more reducing and 
the redox potential of the water decreases. Figure B.3.3 shows the typical ORP conditions for 
ground water when different electron acceptors are used. The main force driving this change in ORP 
is microbially mediated oxidation-reduction reactions. ORP can be used as a crude indicator of 
which oxidation-reduction reactions may be operating at a site. The ORP determined in the field 
using an electrode is termed Eh. Eh can be expressed as pE, which is the hypothetical measure of the 
electron activity associated with a specific Eh. High pE means that the solution or redox couple has 
a relatively high oxidizing potential. 

B.3.2 BIODEGRADATION OF ORGANIC COMPOUNDS VIA USE AS A PRIMARY 
GROWTH SUBSTRATE 

Many organic compounds including natural organic carbon, fuel hydrocarbons, and the less 
oxidized chlorinated compounds such as DCE, 1,2-DCA, chlorobenzene, or vinyl chloride can be 
used as primary growth substrates (electron donor) for microbial metabolism. The following sec¬ 
tions describe biodegradation of organic compounds through use as a primary substrate under both 
aerobic and anaerobic conditions. 

B.3.2.1 Aerobic Biodegradation of Primary Substrates 

Biodegradation of organic compounds is often an aerobic process that occurs when indigenous 
populations of microorganisms are supplied with the oxygen and nutrients necessary to utilize 


B3-33 


Table B.3.3 Trends in Contaminant, Electron Acceptor, Metabolic By-product and Total Alkalinity 
Concentrations During Biodegradation 


Analyte 

Terminal Electron Accepting Process 

Trend in Analyte Concentration During 
Biodegradation 

Fuel Hydrocarbons 

Aerobic Respiration, Denitrification, 
Manganese (IV) Reduction, Iron (III) Reduction, 
Methanogenesis 

Decreases 

Highly Chlorinated Solvents and 
Daughter Products 

Reductive Dechlorination 

Parent Compound Concentration Decreases, Daughter 
Products Increase Initially and Then 

May Decrease 

Lightly Chlorinated Solvents 

Aerobic Respiration, Denitrification, 
Manganese (IV) Reduction, Iron (III) Reduction 
(Direct Oxidation) 

Compound Concentration Decreases 

Dissolved Oxygen 

Aerobic Respiration 

Decreases 

Nitrate 

Denitrification 

Decreases 

Manganese (II) 

Manganese (IV) Reduction 

Increases 

Iron (II) 

Iron (III) Reduction 

Increases 

Sulfate 

Sulfate Reduction 

Decreases 

Methane 

Methanogenesis 

Increases 

Chloride 

Reductive Dechlorination or Direct Oxidation of 
Chlorinated Compound 

Increases 

ORP 

Aerobic Respiration, Denitrification, 
Manganese (IV) Reduction, Iron (III) Reduction, 
Methanogenesis 

Decreases 

Alkalinity 

Aerobic Respiration, Denitrification, Iron (HI) 
Reduction, and Sulfate Reduction 

Increases 


Redox Potential (Eh*) 
In Millivolts @ pH -1 
aridT=25T 


1000 -r 


Aerobic 


Possible Range 
for Reductive 
Dechlorination 


Anaerobic 

500 


o -- 


Optimal Range Tit 
for Reductive 
\ £ Dechlorination 


-500 


O, + 4H* + 4e - > 2H.0 (E„*« ♦ 820) 

2NO,' + 12H**10er-» N. + 6H.0 (V = + 740) 


MnO,(s) + HCO, ♦ 3H* + 2e -> MnCO,(s) * 2H,0 

(E/- + 520) 


FeOOH(s) + HCO, +2W + S 


SO,*> 9K + 8e' 
CO,-*- 8H* + 8e‘ 


FeCO, + 2H,0 
<E»*“-50) 

HS" ♦ 41-1,0 (E,* = - 220) 

CH.-*2H,0 (E»* = - 240) 


Modified From Bouwvr 


(1994) 


Figure B.3.3 Oxidation-reduction potentials for various oxidation-reduction reactions. 


B3-34 























organic carbon as an energy source. The biodegradation of fuel hydrocarbons occurs rapidly under 
aerobic conditions and is discussed in Wiedemeier et al. (1995a). Some pollutants, especially the 
highly oxidized chlorinated hydrocarbons (i.e., those containing more chlorine substituents), are 
biologically recalcitrant under aerobic conditions. However, some of the less chlorinated ethenes 
and ethanes such as DCE, VC, and 1,2-DCA, and many of the chlorinated benzenes can be utilized 
as primary substrates and oxidized under aerobic conditions. During aerobic biodegradation (oxida¬ 
tion) of chlorinated solvents, the facilitating microorganism obtains energy and organic carbon from 
the degraded solvent. 

Of the chlorinated ethenes, vinyl chloride is the most susceptible to aerobic biodegradation, and 
PCE the least. Of the chlorinated ethanes, 1,2-DCA is the most susceptible to aerobic biodegrada¬ 
tion (chloroethane is more likely to abiotically hydrolyze to ethanol), while TCA, tetrachloroethane, 
and hexachloroethane are less so. Chlorinated benzenes with up to 4 chlorine atoms (i.e., chloroben¬ 
zene, dichlorobenzene, trichlorobenzene, and tetrachlorobenzene) also have been shown to be readily 
biodegradable under aerobic conditions (Spain, 1996). Pentachlorobenzene and hexachlorobenzene 
are unlikely to be oxidized by microbial activity. 

B.3.2.1.1 Aerobic Oxidation of Petroleum Hydrocarbons 

Fuel hydrocarbons are rapidly biodegraded when they are utilized as the primary electron donor 
for microbial metabolism under aerobic conditions. Biodegradation of fuel hydrocarbons occurs 
naturally when sufficient oxygen (or other electron acceptors) and nutrients are available in the 
ground water. The rate of natural biodegradation is generally limited by the lack of oxygen or other 
electron acceptors rather than by the lack of nutrients such as nitrogen or phosphorus. The rate of 
natural aerobic biodegradation in unsaturated soil and shallow aquifers is largely dependent upon the 
rate at which oxygen enters the contaminated media. Biodegradation of fuel hydrocarbons is dis¬ 
cussed by Wiedemeier et al. (1995a). 

B.3.2.1.2 Aerobic Oxidation of Chlorinated Ethenes 

In general, the highly chlorinated ethenes (e.g., PCE and TCE) are not likely to serve as electron 
donors or substrates for microbial degradation reactions. This is because the highly chlorinated 
compounds tend to be much more oxidized than many compounds present in a natural ground-water 
system.. Several microbes or microbial enrichments have been shown to be capable of TCE oxida¬ 
tion (Fogel et al ., 1986; Nelson et al., 1986; Little et al., 1988); however, as noted by Vogel (1994), 
no strong evidence for the oxidation of highly chlorinated solvents has been derived from actual 
hazardous waste sites. 

Using microcosms from two different sites with no prior history of exposure to DCE, 

Klier et al. (1998) show that all three isomers of DCE (i.e., 1,1-DCE, I-1,2-DCE, and trans- 1,2- 
DCE) can be biodegraded in aerobic systems. In these experiments, it was observed that cis- 1,2- 
DCE degraded more rapidly than the other isomers. Hartmans et al. (1985) and Hartmans and de 
Bont (1992) show that vinyl chloride can be used as a primary substrate under aerobic conditions, 
with vinyl chloride apparently being directly mineralized to carbon dioxide and water. This has also 
been reported by Davis and Carpenter (1990). Aerobic biodegradation is rapid relative to other 
mechanisms of vinyl chloride degradation, especially reductive dehalogenation. 

B.3.2.1.3 Aerobic Oxidation of Chlorinated Ethanes 

Of the chlorinated ethanes, only 1,2-dichloroethane has been shown to be aerobically mineral¬ 
ized/oxidized. Stucki et al. (1983) and Janssen et al. (1985) show that 1,2-DCA can be used as a 
primary substrate under aerobic conditions. In this case, the bacteria transform 1,2-DCA to 
chloroethanol, which is then mineralized to carbon dioxide. Evidence of oxidation of chloroethane 
is scant, however, it appears to rapidly degrade via abiotic mechanisms (hydrolysis) and is thus less 
likely to undergo biodegradation. 


B3-35 


B.3.2.1.4 Aerobic Oxidation of Chlorobenzenes 

Chlorobenzene and polychlorinated benzenes (up to and including tetrachlorobenzene) have 
been shown to be biodegradable under aerobic conditions. Several studies have shown that bacteria 
are able to utilize chlorobenzene (Reineke and Knackmuss, 1984), 1,4-DCB (Reineke and 
Knackmuss, 1984; Schraaeffl/., 1986; Spain andNishino, 1987), 1,3-DCB (de Bont et al., 1986), 
1,2-DCB (Haigler et al., 1988), 1,2,4-TCB (van derMeer et al, 1987; Sander et al., 1991), and 
1,2,4,5-TeCB (Sander et al ., 1991) as primary growth substrates in aerobic systems. Nishino et al. 
(1994) note that aerobic bacteria able to grow on chlorobenzene have been detected at a variety of 
chlorobenzene-contaminated sites, but not at uncontaminated sites. Spain (1996) notes that this 
provides strong evidence that the bacteria are selected for their ability to derive carbon and energy 
from chlorobenzene degradation in situ. 

The pathways for all of these reactions are similar, and are also similar to that of benzene 
(Chapelle, 1993; Spain, 1996). In general, the aerobic biodegradation involves hydroxylation of the 
chlorinated benzene to a chlorocatechol, followed by ortho cleavage of the benzene ring. This 
produces a muconic acid, which is dechlorinated, and the non-chlorinated intermediates are then 
metabolized. The only significant difference between this process and aerobic benzene degradation 
is the elimination of chlorine at some point in the pathway (Chapelle, 1993). 

B.3.2.2 Anaerobic Biodegradation of Primary Substrates 

Rapid depletion of dissolved oxygen caused by microbial respiration results in the establish¬ 
ment of anaerobic conditions in areas with high organic carbon concentrations. Certain requirements 
must be met in order for anaerobic (anoxic) bacteria to degrade organic compounds, including: 
absence of dissolved oxygen; availability of carbon sources (natural or anthropogenic), electron 
acceptors, and essential nutrients; and proper ranges of pH, temperature, salinity, and redox potential. 
When oxygen is absent, nitrate, manganese (IV), iron (III), sulfate, and carbon dioxide can serve as 
terminal electron acceptors during oxidation of organic carbon. While there is a large body of 
evidence for anaerobic mineralization (oxidation) of fuel hydrocarbons, there is very little evidence 
of such transformations involving chlorinated compounds. 

B.3.2.2.1 Anaerobic Oxidation of Petroleum Hydrocarbons 

Biodegradation of fuel hydrocarbons will occur under anaerobic conditions in most, if not all, 
ground-water environments via denitrification, manganese (IV) reduction, iron (III) reduction, sulfate 
reduction, and methanogenesis. Biodegradation of fuel hydrocarbons is discussed by Wiedemeier et> 
al. (1995a), and many primary references are cited therein. 

B.3.2.2.2 Anaerobic Oxidation of Chlorinated Ethenes 

In general, due to the oxidized nature of polychlorinated ethenes, they are unlikely to undergo 
oxidation in groundwater systems. However, Bradley and Chapelle (1996) show that vinyl chloride 
(with only one chlorine substituent) can be directly oxidized to carbon dioxide and water via 
iron (III) reduction. Reduction of vinyl chloride concentrations in microcosms amended with iron 
(III)-EDTA closely matched the production of carbon dioxide. Slight mineralization was also noted 
in unamended microcosms. The rate of this reaction apparently depends on the bioavailability of the 
iron (III). At this time, it is not known if other workers have demonstrated other anaerobic mineral¬ 
ization reactions involving chlorinated ethenes. 

B.3.2.2.3 Anaerobic Oxidation of Chlorinated Ethanes 

During preparation of this protocol, no evidence of anaerobic oxidation of chlorinated ethanes 
was found; this does not necessarily indicate that such reactions have not been described. However, 
the lack of discussion of such transformations in surveys of chlorinated hydrocarbon biodegradation 
(e.g., Vogel et al., 1987; McCarty and Semprini, 1994; Vogel, 1994, Adriaens and Vogel, 1995; 

Spain, 1996) suggests that there has indeed been little, if any, work on this subject. 


B3-36 


B.3.2.2.4 Anaerobic Oxidation of Chlorobenzenes 

While aerobic mineralization of chlorobenzenes is similar to that of benzene, similar activity 
under anaerobic conditions has not been documented. As discussed above, there is little, if any, 
discussion of this topic in the literature. 

B.3.3 BIODEGRADATION OF ORGANIC COMPOUNDS VIA USE AS AN ELECTRON 
ACCEPTOR (REDUCTIVE DECHLORINATION) 

Bouwer et al. (1981) were the first to show that halogenated aliphatic hydrocarbons could be 
biologically transformed under anaerobic conditions in the subsurface environment. Since that time, 
numerous investigators have shown that chlorinated compounds can degrade via reductive dechlori¬ 
nation under anaerobic conditions. Anaerobically, biodegradation of chlorinated solvents most often 
proceeds through a process called reductive dechlorination. During this process, the halogenated 
hydrocarbon is used as an electron acceptor, not as a source of carbon, and a halogen atom is re¬ 
moved and replaced with a hydrogen atom. As an example, Dehalobacter restrictus was shown by 
Holliger et al., (1993) to use tetrachloroethene as an electron acceptor during reductive dechlorina¬ 
tion to produce c7s-7,2-dichloroethene. Because chlorinated compounds are used as electron accep¬ 
tors during reductive dechlorination, there must be an appropriate source of carbon for microbial 
growth in order for reductive dehalogenation to occur (Baek and Jaffe, 1989; Freedman and Gossett, 
1989; Fathepure and Boyd, 1988; Bouwer, 1994)/ Potential carbon sources can include low molecu¬ 
lar weight organic compounds (lactate, acetate, methanol, glucose, etc.), fuel hydrocarbons, 
byproducts of fuel degradation (e.g., volatile fatty acids), or naturally occurring organic matter. 

In some situations, reductive dechlorination may be a cometabolic process, in that the reaction 
is incidental to normal metabolic functions and the organisms derive no benefit from the reaction. 
Such cometabolism typically results in slow, incomplete dechlorination (Gantzer and Wackett, 1991; 
Gossett and Zinder, 1996). More important, recent studies are discovering direct dechlorinators 
(typically isolated from contaminated subsurface environments or treatment systems) that use chlori¬ 
nated ethenes as electron acceptors in reactions that provide growth and energy (e.g., Holliger et al ., 
1992; Holliger etal. , 1993; Holliger and Schumacher, 1994; Neumann etal., 1994; Krumholz, 1995; 
Maymo-Gatell et al., 1995; Sharma and McCarty, 1996; Gerritse et al., 1996). This process has been 
termed both halorespiration and dehalorespiration. 

Biotic transformations of chlorinated solvents under anaerobic conditions generally are reduc¬ 
tions that involve either hydrogenolysis or dihaloelimination (McCarty and Semprini, 1994). 
Hydrogenolysis occurs when a chlorine atom is replaced with hydrogen. Dihaloelimination occurs 
when two adjacent chlorine atoms are removed and a double bond is formed between the respective 
carbon atoms. The most important process for the natural biodegradation of the more highly chlori¬ 
nated solvents is reductive dechlorination (hydrogenolysis). 

Higher ratios of chlorine to carbon represent higher oxidation levels; highly chlorinated com¬ 
pounds are more oxidized than lesser chlorinated compounds and thus are less susceptible to oxida¬ 
tion. Thus, highly chlorinated compounds such as PCE, TCE, TCA, or HCB are more likely to 
undergo reductive reactions than oxidative reactions. During these reductive reactions, electrons are 
transferred to the chlorinated compound, and a chlorine atom is replaced with a hydrogen atom. As 
an example, consider the reductive dechlorination of PCE to TCE and then TCE to DCE, and finally 
DCE to vinyl chloride. Because of the relatively low oxidation state of VC, this compound more 
commonly undergoes aerobic biodegradation as a primary substrate than reductive dechlorination. 

Reductive dechlorination processes result in the formation of intermediates which are more 
reduced than the parent compound. These intermediates are often more susceptible to oxidative 
bacterial metabolism than to further reductive anaerobic processes. Actual mechanisms of reductive 
dehalogenation are still unclear, and in some cases may be a form of cometabolism (Gantzer and 
Wackett, 1991; Adriaens and Vogel, 1995; Wackett, 1995). In addition, other factors that will influ- 


B3-37 


ence the process include the type of electron donor and the presence of competing electron acceptors 
(Adriaens and Vogel, 1995; Suflita and Townsend, 1995), temperature, and substrate availability. 

Recent evidence suggests that dechlorination is dependent upon the supply of hydrogen (H 2 ), 
which acts as the electron donor in many such reactions (Gossett and Zinder, 1996; Smatlak et al., 
1996). The hydrogen is produced as a result of the microbial degradation of a primary substrate 
(e.g., lactate, acetate, butyrate, ethanol, BTEX, or other such compounds). Bacteria that facilitate 
dechlorination compete with sulfate-reducers and methanogens for the H 2 produced in such a system. 
When degradation of the original substrate/electron donor rapidly yields high concentrations of H 2 , 
the sulfate-reducers and methanogens appear to be favored over the dechlorinators. Conversely, 
when substrate degradation produces a steady supply of H 2 at low concentrations, the dechlorinators 
are favored (Gossett and Zinder, 1996; Smatlak et al ., 1996). Complete dechlorination is thus 
apparently favored when a steady, low-concentration supply of H 2 is produced through microbial 
degradation of substrates such as proprionate or benzoate (and, by extension from benzoate, the 
BTEX compounds) (Gossett and Zinder, 1996). Therefore, the type of substrate/electron donor can 
also play a role in how thoroughly a natural system is able to dechlorinate solvents. 

One or more of the following generally is observed at a site where reductive dechlorination of 
alkenes is ongoing: 

1) Ethene is being produced (even low concentrations are indicative of biodegradation); 

2) Methane is being produced; 

3) Iron II is being produced; 

4) Hydrogen concentrations are between 1-4 nM; and 

5) Dissolved oxygen concentrations are low. 

B.3.3.1 Reductive Dechlorination of Chlorinated Ethenes 

PCE and TCE have been shown to undergo reductive dechlorination in a variety of anaerobic 
systems from different environments, with various electron donors/carbon sources (Table B.3.4) 
(Wilson, 1988; Sewell et al., 1991; Roberts et al., 1982). This is particularly true if the subsurface 
also contains other anthropogenic or native organic compounds that can serve as electron donors and 
whose utilization by subsurface bacteria will deplete any available oxygen. In general, reductive 
dechlorination of chlorinated ethenes occurs by sequential dechlorination from PCE to TCE to DCE 
to VC to ethene. Depending upon environmental conditions, this sequence may be interrupted, with 
other processes then acting upon the products. With sufficient quantities or appropriate types of 
electron donors (e.g., slow but steady H 2 -production), the final end-product of anaerobic reductive 
dehalogenation can be ethene (Freedman and Gossett, 1989). Reductive dehalogenation of chlori¬ 
nated solvent compounds is associated with the accumulation of daughter products and an increase 
in chloride. 

Studies have shown that PCE and TCE can be anaerobically reduced to either 1,1-DCE, cis- 1,2- 
DCE, or trans- 1,2-DCE, all of which can be further transformed to vinyl chloride (Miller and 
Guengerich, 1982; Wilson and Wilson, 1985; Mayer et al ., 1988; Nelson, et al ., 1986; Henson et al ., 
1989; Tsien etal, 1989; Henry, 1991; McCarty, 1994; Wilson etal., 1994). During reductive 
dehalogenation, all three isomers of DCE can theoretically be produced; however, Bouwer (1994) 
reports that cis-l,2-DCE is a more common intermediate than trans-l,2-DCE and that 7,7-DCE is 
the least prevalent intermediate of the three DCE isomers. Vinyl chloride produced from 
dehalogenation of DCE may be subsequently reduced to innocuous products such as ethane or 
carbon dioxide. The removal of vinyl chloride occurs more readily under aerobic conditions, such as 
those encountered at the edge of the plume. Vinyl chloride may also be used as a primary substrate 
by aerobic organisms, as previously discussed. 


B3-38 


Table B.3.4 Sources, Donors, Acceptors, and Products of Reductive Dechlorinating Laboratory Systems 


Reference 

Source 

Donor 

Acceptor-Product 

Bouwer & McCarty, 1983 

Digester 

Organic Material 

PCE-TCE 

Vogel & McCarty, 1985 

Bioreactor 

Acetate 

PCE-VC, C0 2 

Kleopfer etal., 1985 

Soil 

Soybean Meal 

TCE-DCE 

Barrio-Lage etal., 1987 

Swamp Muck 

Organic Material 

PCE-VC 


Soil 

Methanol (?) 

PCE-VC 

Fathepure etal., 1987 

Methanosarcina 

Methanol 

PCE-TCE 


DCB-1 

3CB a ,Pyruvate,RF b 

PCE-TCE 

Back & Jaffe, 1989 

Digester 

Formate 

TCE-VC,CA C 



Methanol 

TCE-VC,CA 

Freedman & Gossett, 1989 

Digester 

Methanol 

PCE-VC, Ethene 



Glucose 

PCE-VC, Ethene 



H2 

PCE-VC, Ethene 



Formate 

PCE-VC, Ethene 



Acetate 

PCE-VC, Ethene 

Scholz-Muramatsu et al., 1990 

Bioreactor 

Benzoate 

PCE-DCE 

Gibson & Sewell, 1990 

Aquifer 

VFA d 

PCE-DCE 

Sewell & Gibson, 1990 

Aquifer 

Toluene 

PCE-DCE 

Sewell etal., 1991 

Aquifer 

VFA 

PCE-DCE 


Landfill 

VFA 

PCE-VC 

Lyon et al., 1995 

Aquifer 

Native Organic Matter 

PCE-DCE 


a 3-Chlorobenzoate 
b Rumen Fluid 
c Chloroethane 
d Volatile Fatty Acid 


B3-39 



















B.3.3.2 Reductive Dechlorination of Chlorinated Ethanes 

As with the ethenes, chlorinated ethanes will also undergo reductive dehalogenation in the 
subsurface via use as electron acceptors. Dechlorination of TCA has been described by Vogel and 
McCarty (1987) and Cox et al. (1995), but this pathway is complicated by the abiotic reactions that 
can affect TCA and its byproducts (Vogel, 1994). 

B.3.3.3 Reductive Dechlorination of Chlorobenzenes 

For the highly chlorinated benzenes (e.g., hexachlorobenzene and pentachlorobenzene, as well 
as tetrachlorobenzene, and trichlorobenzene), reductive dechlorination is the most likely biodegrada¬ 
tion mechanism (Holliger et al., 1992; Ramanand et al., 1993; Suflita and Townsend, 1995). As 
discussed by Suflita and Townsend (1995), reductive dehalogenation of aromatic compounds has 
been observed in a variety of anaerobic habitats, including aquifer materials, marine and freshwater 
sediments, sewage sludges, and soil samples; however, isolation of specific microbes capable of 
these reactions has been difficult. As with the chlorinated ethenes and ethanes, the chlorobenzenes 
are most likely acting as electron acceptors as other sources of carbon and energy are being utilized 
by microbes or microbial consortia (Suflita and Townsend, 1995). Evidence has been presented 
suggesting that oxidation of hydrogen using halogenated aromatics as electron acceptors may yield 
more energy than if more commonly available electron acceptors were used (Dolfing and Harrison, 
1992). 

As discussed previously, the actual mechanisms of reductive dehalogenation are not well under¬ 
stood. Further, reductive dehalogenation of chlorinated benzenes has not been as well-documented 
as for other chlorinated solvents. However, reductive dechlorination of chlorobenzenes has been 
documented more frequently in the past several years (e.g., Bosma et al., 1988; Fathepure et al., 

1988; Fathepure and Vogel, 1991; Holliger et al., 1992; Ramanand et al., 1993). As with other 
chlorinated solvents, the reductive dehalogenation of chlorobenzenes is affected by the degree of 
chlorination of the compound. The more chlorinated aromatic compounds are typically more ame¬ 
nable to this reaction (Suflita and Townsend, 1995; Adriaens and Vogel, 1995), but as they are 
dechlorinated, the daughter products will become more resistant to further dehalogenation reactions 
(Fathepure et al., 1988; Bosma et al., 1988; Holliger et al., 1992). The reductive dechlorination of 
chlorobenzenes is analogous to reactions involving chlorinated ethenes and ethanes in that such 
degradation will make them more amenable to aerobic biodegradation (Schraa, et al., 1986; Spain 
and Nishino, 1987; Ramanand et al., 1993). , 

B.3.4 BIODEGRADATION OF ORGANIC COMPOUNDS VIA COMETABOLISM 

When a chlorinated solvent is biodegraded through cometabolism, it serves as neither an elec¬ 
tron acceptor nor a primary substrate in a biologically mediated redox reaction. Instead, the degrada¬ 
tion of the compound is catalyzed by an enzyme cofactor that is fortuitously produced by organisms 
for other purposes. The best-documented cometabolism reactions involve catabolic oxygenases that 
catalyze the initial step in oxidation of their respective primary or growth substrate (BTEX or other 
organic compounds). These oxygenases are typically nonspecific and, therefore, fortuitously initiate 
oxidation of a variety of compounds, including many of the CAHs (McCarty and Semprini, 1994). 
The organism receives no known benefit from the degradation of the chlorinated solvent; in some 
cases the cometabolic degradation of the solvent may, in fact, be harmful to the microorganism 
responsible for the production of the enzyme or cofactor (McCarty and Semprini, 1994). Chlorinated 
solvents are usually only partially transformed during cometabolic processes, with additional biotic 
or abiotic degradation generally required to complete the transformation (McCarty and Semprini, 
1994). 

Cometabolism is best documented for CAHs in aerobic environments; evidence of 
cometabolism of chlorobenzenes is scant, as is clear evidence of anaerobic cometabolism. In an 


B3-40 


aerobic environment, many chlorinated organic compounds can only be degraded via cometabolism. 
It has been reported that under aerobic conditions chlorinated ethenes, with the exception of PCE, 
are susceptible to cometabolic degradation (Murray and Richardson, 1993; Vogel, 1994; McCarty 
and Semprini, 1994; Adriaens and Vogel, 1995). Vogel (1994) further elaborates that the oxidation 
rate increases as the degree of chlorination decreases. Aerobic cometabolism of ethenes may be 
characterized by a loss of contaminant mass, the presence of intermediate degradation products (e.g., 
chlorinated oxides, aldehydes, ethanols, and epoxides), and the presence of other products such as 
chloride, carbon dioxide, carbon monoxide, and a variety of organic acids (Miller and Guengerich, 
1982; McCarty and Semprini, 1994). 

The lack of clear evidence for anaerobic cometabolism does not necessarily imply that such 
transformations do not occur; in some cases, reductive dechlorination may be a result of 
cometabolism (e.g., Gantzer and Wackett, 1991), depending upon the relationship between the 
microbes, substrates, contaminants, and other electron acceptors. However, as with aerobic 
cometabolism, anaerobic cometabolism will be slow relative to dehalorespiration and might not be 
distinguishable at the field scale (Gossett and Zinder, 1996). 

Several groups of aerobic bacteria currently are recognized as being capable of transforming 
TCE and other CAHs via cometabolism; these groups include: 

• Methane Oxidizers (Methanotrophs) (Fogel et al., 1986; Little et al ., 1988, Mayer et al., 
1988; Oldenhuis et al., 1989; Tsien et al., 1989; Henry and Grbic-Galic, 1990; Alvarez- 
Cohen and McCarty, 1991a,b; Henry and Grbic-Galic, 1991a,b; Lanzarone and McCarty, 
1990; Oldenhuis etal., 1991); 

• Propane Oxidizers (Wackett et al., 1989); 

• Ethene Oxidizers (Henry, 1991); 

• Toluene, Phenol, or Cresol Oxidizers (Nelson et al., 1986,1987, 1988; Wackett and 
Gibson, 1988; Folsom etal., 1990; Harker and Kim, 1990); 

• Ammonia Oxidizers (Arciero et al., 1989; Vannelli et al., 1990); 

• Isoprene Oxidizers (Ewers et al., 1991); and 

• Vinyl Chloride Oxidizers (Hartmans and de Bont, 1992). 

These bacteria all have catabolic oxygenases that catalyze the initial step in oxidation of their 
respective primary or growth substrates and have the potential for initiating the oxidation of CAHs. 

Cometabolism is not nearly as important a degradation mechanism for chlorinated solvents in 
the saturated zone as reductive dehalogenation. Due to the need for a substrate that may be present 
in limited concentrations, as well as the fortuitous nature of the reactions, rates of cometabolism are 
often slow enough that this process may not be detectable unless the system is stimulated with 
additional substrate mass. For a discussion of this topic, see McCarty and Semprini (1994) or 
Wackett (1995). 

B.3.5 THERMODYNAMIC CONSIDERATIONS 

Electron transfer results in oxidation of the electron donor and reduction of the electron accep¬ 
tor and the production of usable energy. The energy produced by these reactions is quantified by the 
Gibbs free energy of the reaction (G) which is given by: 

AG; = X -1A G}„ eq. B.3.1 

Where: 

AG r = Gibbs Free Energy of the Reaction at Standard State 
A G f producB = Gibbs Free Energy of Formation for Products at Standard State 
A Gy reacuutts = Gibbs Free Energy of Formation for the Reactants at Standard State 


B3-41 


The G r defines the maximum useful energy change for a chemical reaction at a constant tem¬ 
perature and pressure. Table B.3.5 presents select electron acceptor and electron donor half-cell 
reactions and the calculated G r values. Table B.3.6 gives the Gibbs free energy of formation (G f ) for 
species used in these half-cell reactions. Table B.3.7 presents coupled oxidation-reduction reactions. 
In general, those reactions that yield the most energy tend to take precedence over less energy- 
yielding reactions. However, the calculated energy yield of processes involving anthropogenic 
organic compounds may not be reflected in the true energy yield of the metabolic process. 

Figure B.3.4 illustrates the expected sequence of microbially mediated redox reactions based on G r . 
There is sufficient energy in the reaction of fuel hydrocarbons with chlorinated solvents to allow 
their use by microorganisms as physiological electron acceptors. 


B3-42 


Table B.3.5 Electron Donor and Electron Acceptor Half-Cell Reactions 


HALF-CELL REACTIONS 

AG°,(kcal/ 

equiv)’ 

AG°,(kJ/ 

equiv)* 

E° 

(V) 

Eh 

(V) 

pe 

Conditions 
for Eh and pe § 

ELECTRON-ACCEPTOR (REDUCTION) HALF CEL 

X REACTION 

S 





5e + 6tT + NOj => 0.5N 2 + 3H 2 0 

Denitrification 

-28.7 

-120. 

+ 1.24 

+0.708 

+12.0 

pH = 7 
irNi=io- 3 

4e + 4H* + 0 2 => 2H 2 0 

Aerobic Respiration 

-28.3 

-119. 

+ 1.23 

+0.805 

+13.6 

pH = 7 

Po =0.21 atm 

2 

2e + 4 + MnO ? => Mn 2 * + 2H 2 0 

Pyrolusite Dissolution/Reduction 

-28.3 

-119 

+1.23 

+1.169 

+19.8 

pH = 7 
I[Mnl=10- 5 

CO, + e+ H* + MnOOH => MnCO , + H,0 

Manganite Carbonation/Reduction 

-23.1 

-96.8 

+1.00 

+0.408 

+6.90 

pH = 8 

Pco =10' 2 

2 

e + FT + MnO,=> MnOOH 

Pyrolusite Hydrolysis/Reduction 

-22.1 

-92.5 

+0.959 

+0.545 

+9.21 

pH = 7 

e + 3H* + Fe(OH ^ Fe 2+ + 

Amorphous "Goethite " Dissolution/Reduction 

-21.5 

-89.9 

+0.932 

+0.163 

+2.75 

pH = 6 
irFel=10' 5 

8e + 79/T + AO, => NH* 4 + i//,0 

Nitrate Reduction 

-20.3 

-84.9 

+0.879 

+0.362 

+6.12 

pH = 7 

2e +2H* + NO'} => AO', + A,0 

Nitrate Reduction 

-18.9 

-78.9 

+0.819 

+0.404 

+6.82 

pH = 7 

e + 3H J + FeOOH => + 2H,0 

"Ferric oxyhydroxide" Dissolution/Reduction 

-15.0 

-62.9 

+0.652 

-0.118 

-1.99 

pH = 6 
£ fFel=10' s 

e' + 3H + + Fe(OH => Fe" + + iA,0 

Crystallized "Goethite" Dissolution/Reduction 

-11.8 

-49.2 

+0.510 

-0.259 

-4.38 

pH = 6 
£ [Fe]=10" s 

e' + /A + CO>„ + Fe(OH )\=> FeCOt + 2H->0 
Amorphous "Goethite" Carbonation/Reduction 

-11.0 

-46.2 

+0.479 

-0.113 

-1.90 

pH = 8 

Pco, =10‘ 2 atm 

8e + 9FT + SO 2 , =>AA + 4H 2 0 

Sulfate Reduction 

-5.74 

-24.0 

+0.249 

-0.278 

-4.70 

pH = 8 

8e + 10H* + SO 2 , => + 4H 2 0 

Sulfate Reduction 

-6.93 

-28.9 

+0.301 

-0.143 

-2.42 

pH = 6 

8e + 8H* + C0 2 , g => CH 4g + 2H 2 0 

Methanogenesis 

-3.91 

-16.4 

+0.169 

-0.259 

-4.39 

pH = 7 

Pco 2 =10- 2 

Pch 4 = 10° 

C 2 Cl 4 + FT + 2e => C 2 HCl} + Cf 

PCE Reductive Dechlorination 

-14.79 

-61.8 

+0.641 

+0.552 

+9.33 

pH = 7 
[Cl-]=10' 4 

C 2 HCl 3 + /T + 2e=> C 2 H 2 Cl 2 + ci 

TCE Reductive Dechlorination 

-14.50 

-60.6 

+0.628 

+0.539 

+9.12 

pH = 7 
[Cl-]=10' 4 

C 2 H 2 Cl 2 + FT + 2e' => C 2 H 3 Cl + Cf 
c-DCE Reductive Dechlorination 

-12.12 

-50.7 

+0.525 

+0.436 

+7.38 

pH = 7 
[Cl-]=10' 4 

C 2 H}Cl + H* + 2e=> C 2 H 4 + CL 

FC Reductive Dechlorination 

-13.75 

-57.5 

+0.596 

+0.507 

+8.57 

pH = 7 

rci-i=io- 4 

C 2 H 2 Cl 4 + /T + 2e => C 2 H}Cl} + CL 

PC4 Reductive Dechlorination 

-13.59 

-56.8 

+0.589 

+0.500 

+8.45 

pH = 7 
rci-Mo- 4 

C 2 H}Cl } +FT + 2e=> C 2 H 4 Cl 2 + CL 

TC4 Reductive Dechlorination 

-15.26 

-63.8 

+0.661 

+0.572 

+9.67 

pH = 7 
[Cl-l=10^ 

C 2 H£l 2 + EF + 2e=> C 2 H s Cl + Cl 

DCA Reductive Dechlorination 

-14.08 

-58.9 

+0.610 

+0.521 

+8.81 

pH = 7 

rci-wo- 4 

C 6 Cl 6 +H* + 2e => CtHCls + Cl 
Hexachlorobenzene Reductive Dechlorination 

-14.36 

-60.0 

+0.622 

+0.533 

+9.01 

pH = 7 

rci-wo- 4 

CtHCls + H* + 2e' => CfH 2 Cl 4 + Cl 
Pentachlorobenzene Reductive Dechlorination 

-14.64 

-61.2 

+0.634 

+0.545 

+9.22 

pH = 7 

rci-Mo- 4 

C<H 2 Cl 4 + H* + 2e CeHjCh + Cl 
Tetrachlorobenzene Reductive Dechlorination 

-13.66 

-57.1 

+0.592 

+0.503 

+8.50 

pH = 7 

rci-Mo- 4 

C«H}Cl } +H* + 2e => C<H 4 Cl 2 + Cl 
Trichlorobenzene Reductive Dechlorination 

-13.20 

-55.2 

+0.572 

+0.483 

+8.17 

pH = 7 

[Cl-]=10‘ 4 


B3-43 




























































Table B. 3.5 Continued. 


HALF-CELL REACTIONS 

AG° r (kcal/ 

equiv)’ 

AG° r (kJ/ 

equiv)’ 

E° 

(V) 

Eh 

(V) 

pe 

Conditions 
for Eh and pe § 

ELECTRON-DONOR (OXIDATION) HALF CELL REACTIONS 







12HJ0 + CJi 6 => 6C0 2 + 30W + 30e 

Benzene Oxidation 

+2.83 

+11.8 

-0.122 

+0.316 

+5.34 

pH = 7 
Pco 2 =10‘ 2 

14H 2 0 + CsHsCHs => 7C0 2 + 36FT + 36e 

Toluene Oxidation 

+2.96 

+12.4 

-0.128 

+0.309 

+5.22 

pH = 7 
Pco 2 =10- 2 

16H 2 0 + CsHsCiHs => 8C0 2 + 42TT + 42e 

Ethylbenzene Oxidation 

+2.96 

+12.4 

-0.128 

+0.309 

+5.21 

pH = 7 
Pco 2 =10- 2 

16H 2 0 + CtHJCH 3 ) 2 => 8C0 2 + 42W + 42e 
m-Xylene Oxidation 

+3.03 

+12.7 

-0.132 

+0.303 

+5.12 

pH = 7 
Pco 2 =10- 2 

20H 2 O + CjoHb => 10CO 2 + 48W + 48e 

Naphthalene Oxidation 

+2.98 

+12.5 

-0.130“ 

+0.309 

+5.22 

pH = 7 

Pco 2 =10‘ 2 

18H 2 0 + C«H 3 (CH })3 => 9C0 2 + 48H+ + 48e 

1,3,5-Trimethylbenzene Oxidation 

+3.07 

+12.8 

-0.133“ 

+0.303 

+5.12 

pH = 7 
Pco 2 =10- 2 

18H 2 0 + CtHitCHji => 9C0 2 + 48H+ + 48e 
1,2,4-Trimethylbenzene Oxidation 

+3.07 

+12.9 

-0.134“ 

+0.302 

+5.11 

pH = 7 
Pco 2 = 10‘ 2 

4H 2 0 + C 2 H 2 Cl 2 => 2C0 2 + 10W + 8e + 2CT 

DCE Oxidation 

-3.88 

-16.2 

+0.168 

-0.131 

-2.21 

pH = 7 
Pco 2 =10- 2 

4H 2 0 + C 2 H 3 Cl => 2C0 2 + 1 nr + IOe + Cl 

Vinyl Chloride Oxidation 

-0.55 

-2.31 

+0.024“ 

-0.006 

-0.10 

pH = 7 
Pco 2 =10- 2 

12H 2 0 + C<n 2 Cl 4 =o 6C0 2 + 26W + 22e + 4CT 
Tetrachlorobenzene Oxidation 

-0.64 

-2.68 

+0.028 

+0. 016 

+0.27 

pH = 7 
Pco 2 =10' 2 

12H 2 0 + CeHiCh => 6C0 2 + 27TT + 24e + 3CI 
Trichlorobenzene Oxidation 

+0.42 

+1.77 

-0.018 

-0.030 

-0.50 

pH = 7 
Pco 2 =10- 2 

12H 2 0 + C<W 4 Cl 2 => 6C0 2 + 28H* + 26e + 2CT 
Dichlorobenzene Oxidation 

+1.40 

+5.84 

-0.060 

-0.071 

-1.21 

pH = 7 

Pco 2 = 10' 2 

12H 2 0 + CMsCl => 6C0 2 + 29W + 28e + Cl 
Chlorobenzene Oxidation 

+2.22 

+9.26 

-0.096“ 

-0.107 

-1.80 

pH = 7 

Pco 2 =10- 2 


NOTES: 

* = AG° r for half-cell reaction as shown divided by the number of electrons involved in reaction. 

§ = Conditions assumed for the calculation of Eh and pe (pe = Eh/0.05916). Where two dissolved species are involved, 
other than those mentioned in this column, their activities are taken as equal. Note, this does not affect the free 
energy values listed. 

8 = E° calculated using the following equation; E° = AG° r (J/nF) * 1.0365x1 O' 5 (VF/J) from Stumm and Morgan, 1981. 


B3-44 







































Table B.3.6 Gibbs Free Energy of Formation for Species used in Half-Cell Reactions and Coupled 

Oxidation-Reduction Reactions 


Species 

State 

AG°f,298.15 

(kcal/mole) 

Source 

e' 

i 

0 

std 

H + 

i 

0 

std 

0 2 

g 

0 

std 

H 2 0 

1 

-56.687 

Dean (1972) 

Carbon Species 

C0 2 

g 

-94.26 

Dean (1972) 

CH 2 0, formaldehyde 

aq 

-31.02 

Dean (1972) 

C 6 H 6 , benzene 

1 

+29.72 

Dean (1972) 

CH 4 , methane 

g 

-12.15 

Dean (1972) 

C 6 H 5 CH 3 , toluene 

1 

+27.19 

Dean (1972) 

C 6 H 5 C 2 H 5 , ethylbenzene 

1 

+28.61 

Dean (1972) 

C 6 H 4 (CH 3 ) 2 , o-xylene 

1 

+26.37 

Dean (1972) 

C 6 H 4 (CH 3 ) 2 , m-xylene 

1 

+25.73 

Dean (1972) 

C 6 H 4 (CH 3 ) 2 , p-xylene 

1 

+26.31 

Dean (1972) 

CjCU, PCE 

1 

+1.1 

CRC Handbook (1996) 

C 2 HC1 3 , tce 

1 

+2.9 

CRC Handbook (1996) 

C 2 H 2 C 1 2 1,1-dichloroethene 

1 

+5.85 

Dean (1972) 

C 2 H 2 C 1 2 cis-1,2-dichloroethene 

1 

5.27 

CRC Handbook (1996) 

C 2 H 2 C 1 2 trans-1,2- 
dichloroethene 

1 

+6.52 

CRC Handbook (1996) 

QH 4 Ethene 

g 

aq,m=l 

+16.28 

+19.43 

CRC Handbook (1996) 

C 2 H 6 Ethane 

g 

aq, m=l 

-7.68 

-4.09 

CRC Handbook (1996) 

HC1 hydrochloric acid 

aq, m=l 

-31.372 

CRC Handbook (1996)a 

C 2 H 2 C1 4 ,1,1,2,2-PCA 

1 

-22.73 

Dean (1972) 

C^Ch, 1,1,2-TCA 

g 

-18.54 

Dean (1972) 

CACb, 1,2-DCA 

g 

-17.68 

Dean (1972) 

QHsCli, Chloroethane 

g 

-14.47 

Dean (1972) 

CioHg, naphthalene 

1 

+48.05 

Dean (1972) 

CfiH 3 (CH 3 ) 3 , 1,3,5-TMB 

1 

+24.83 

Dean (1972) 

CfiH 3 (CH 3 ) 3 , 1,2,4-TMB 

1 

+24.46 

Dean (1972) 

C 2 H 3 C1, Vinyl chloride 

g 

+12.4 

Dean (1972) 

C 6 C1 6 , Hexachlorobenzene 

1 

+0.502 

Dotting and Harrison (1992) 

C 6 HiC 1 5 , Pentachlorobenzene 

1 

+3.16 

Dotting and Harrison (1992) 

C 6 H 2 C1 4 ,1,2,4,5- 
Tetrachlorobenzene 

1 

+5.26 

Dotting and Harrison (1992) 

C 6 H 3 C1 3 ,1,2,4- 
Trichlorobenzene 

1 

+9.31 

Dolfing and Harrison (1992) 

C 6 H 4 C1 2 , 1,4-Dichlorobenzene 

1 

+ 14.28 

Dotting and Harrison (1992) 

CfiHsCl, chlorobenzene 

1 

+21.32 

Dean (1972) 

Ci 4 Hio, phenanthrene 

1 

+64.12 

Dean (1972) 


B3-45 
























































Table B. 3.6 Continued. 


Species 

State 

A(j°f >2?8 .i5 

(kcal/mole) 

Source 

Nitrogen Species 

no 3 

I 

-26.61 

Dean (1972) 

n 2 

g 

0 

std 

N0 2 * 

I 

-7.7 

Dean (1972) 

nh 4 + 

aq 

-18.97 

Dean (1972) 

Sulfur Species 

S0 4 2 - 

i 

-177.97 

Dean (1972) 

h 2 s 

aq 

-6.66 

Dean (1972) 

h 2 s 

g 

-7.9 

Dean (1972) 

HS' 

i 

+2.88 

Dean (1972) 

Iron Species 

Fe 2+ 

i 

-18.85 

Dean (1972) 

F? 7 

i 

-1.1 

Dean (1972) 

Fe 2 0 3 , hematite 

c 

-177.4 

Dean (1972) 

FeOOH, ferric oxyhydroxide 

c 

-117.2 

Naumov et al. (1974) 

Fe(OH) 3 , goethite 

a 

-167.416 

Langmuir and Whittemore 
(1971) 

Fe(OH) 3 , goethite 

c 

-177.148 

Langmuir and Whittemore 
(1971) 

FeC0 3 , siderite 

c 

-159.35 

Dean (1972) 

Manganese Species 

Mn 2+ 

i 

-54.5 

Dean (1972) 

Mn0 2 , pyrolusite 

c 

-111.18 

Stumm and Morgan 
(1981) 

MnOOH, manganite 

c 

-133.29 

Stumm and Morgan 
(1981) 

MnC0 3 , rhodochrosite 

P 

-194 

Dean (1972) 

Chloride Species 

Cl* 

aq 

-31.37 

Dean (1972) 


NOTES: 

c = crystallized solid 1 = liquid g = gaseous aq = undissociated aqueous species 

a = amorphous solid (may be partially crystallized - dependent on methods of preparation) 
p = freshly precipitated solid 

i = dissociated, aqueous ionic species (concentration = 1 m) 
std = accepted by convention 

Wherever possible multiple sources were consulted to eliminate the possibility of typographical error. 


B3-46 



















































Table B.3. 7 Coupled Oxidation-Reduction Reactions 


Coupled Benzene Oxidation Reactions 

AG° r 

(kcal/mole) 

AG° r 

(kJ/ mole) 

Stoichiometric 
Mass Ratio of 
Electron Acceptor 
or Metabolic 
Byproduct to 
Primary Substrate 

Mass of Primary 
Substrate Utilized per 
Mass of Electron 
Acceptor Utilized or 
Metabolic Byproduct 
Produced 

7.50 2 + C 6 H 6 =* 6C0 2 . g + 3H 2 0 

Benzene oxidation /aerobic respiration 

-765.34 

-3202 

3.07:1 

0.326:1 

6NO-3 + 6H+ + C 6 H 6 => 6C0 2 . g + 6H 2 0 + 

Benzene oxidation / denitrification 

-775.75 

-3245 

4.77:1 

0.210:1 

30H + + l5Mn0 2 + C 6 H 6 => + 15Mn 2+ + 7S// 2 0 

Benzene oxidation / manganese reduction 

-765.45 

-3202 

10.56:1 

0.095:1 

3.75 NOj- + C 6 H 6 + 7.5 H + + 0.75 H 2 0 => 6 C0 2 + 3.75 NH 4 + 
Benzene oxidation / nitrate reduction 

-524.1 

-2193 

2.98:1 

0.336:1 

60H* + 30Fe(OH) 3a + C 6 H 6 => 6C0 2 + 30 Fe 2 * + 78H 2 0 

Benzene oxidation / iron reduction 

-560.10 

-2343 

21.5:1 

0.047:1 

75 H + + 3.75 SO If + C 6 H 6 => <SC0 2 , g + 3.15 H 2 S° + 3H 2 0 
Benzene oxidation /sulfate reduction , 

-122.93 

-514.3 

4.61:1 

0.22:1 

4.5H 2 0 + C 6 H 6 => 2.25C0 2ig + 3.75CH 4 

Benzene oxidation / methanogenesis 

-32.40 

-135.6 

0.77:1 

1.30:1 

15 C 2 H 2 C1 4 + C 6 H s + 12 H 2 0 => 6 C0 2 + 15 C 2 H 3 C1 3 +15 H + + 15 Cl' 
Benzene oxidation / PCA reduction 

-322.7 

-1349 

32.2:1 

0.03:1 

15 C 2 H 3 C1 3 + C 6 H 6 + 12 H 2 0 6 => C0 2 + 15 C 2 H 4 C1 2 +15 H + + 15 Cl' 
Benzene oxidation / TCA reduction 

-372.65 

-1558 

25.6:1 

0.04:1 

15 C 2 H 4 C1 2 + CeRe + 12 H 2 0 => 6 C0 2 + 15 C 2 H 5 C1 +15 H + + 15 Cl' 
Benzene oxidation / DCA reduction 

-337.40 

-1410 

19.0:1 

0.05:1 

15C 2 Cl 4 + 12H 2 0 + CiH 6 => 15C 2 HCl s + 6CO : + 15FT + 15CI 
Benzene oxidation/ Tetrachloroethylene reductive dehalogenation 

-358.55 

-1499 

31.8:1 

0.03:1 

15C 2 HCl 3 + 12H 2 0 + CiH 6 => 15C 2 H 2 C1 2 + 6C0 2 + 15Ff + l5Ct 
Benzene oxidation/ Trichloroethylene reductive dehalogenation 

-331.25 

-1385 

25.2:1 

0.04:1 

15C 2 H 2 Cl 2 + 12H 2 0 + CeHe => 15C 2 H 3 Cl + 6CO : + 15fF + l5Ct 
Benzene oxidation/ cis-Dichloroethylene reductive dehalogenation 

-297.35 

-1243 

18.6:1 

0.05:1 

15C 2 H 3 Cl + I2H 2 0 + CsH 6 => 15C 2 H 4 + 6CO : + 15fF + 15CT 
Benzene oxidation/ Vinyl chloride reductive dehalogenation 

-327.35 

-1368 

12.0:1 

0.08:1 

15C 6 Cl 6 + 12H 2 0 + CsHs => 15CfH,Cl s + <JC0 2 + 75/T + l5Ct 

Benzene oxidation/ Hexachlorobenzene reductive dehalogenation 

-345.68 

-1445 

54.7:1 

0.02:1 

ISCeHiCls + 12H 2 0 + => 15C(P 2 C1 4 + 6C0 2 + 75/T + 75CC 

Benzene oxidation/ Pentachlorobenzene reductive dehalogenation 

-354.05 

-1480 

48.1:1 

0.02:1 

15C^i 2 Cl 4 + 12H 2 0 + Cffli => l5CtFl 3 Cl 3 + dC0 2 + 75/T + 15Ct 

Benzene oxidation/ Tetrachlorobenzene reductive dehalogenation 

-324.80 

-1358 

41.5:1 

0.02:1 

ISCsHsCls + 12H 2 0 + C(H 6 => 15CsH 4 Cl 2 + dC0 2 + 15 If + 75C/ 

Benzene oxidation/ Trichlorobenzene reductive dehalogenation 

-311.0 

-1300 

34.8:1 

0.03:1 


B3-47 





































Table B. 3 .7 Continued. 


Coupled Toluene Oxidation Reactions 

AG° r 

(kcal/ mole) 

AG°r 

(kJ/ mole) 

Stoichiometric 
Mass Ratio of 
Electron Acceptor 
or Metabolic 
Byproduct to 
Primary Substrate 

Mass of Primary 
Substrate Utilized per 
Mass of Electron 
Acceptor Utilized or 
Metabolic Byproduct 
Produced 

90 2 + C 6 HsCH 3 => 7CO lg + 4H 2 0 

Toluene oxidation /aerobic respiration 

-913.76 

-3823 

3.13:1 

0.32:1 

7.2 NO) + 7.2/T + CtHiCH) =>7C0 2g + 7.6H 2 0 + 3.6N 2 . g 
Toluene oxidation / denitrification 

-926.31 

-3875 

4.85:1 

0 . 21:1 

36H* + 18MnO> + C 6 H 5 CH 3 => 7 C0 2g + 18Mn 2 + + 22H 2 0 
Toluene oxidation / manganese reduction 

-913.89 

-3824 

10.74:1 

0.09:1 

72H* + 36Fe(OH) }a + C 6 H 5 CH } => 7C0 2 + 36Fe } * + 94H 2 0 

Toluene oxidation / iron reduction 

-667.21 

-2792 

21 . 86:1 

0.05:1 

9H* + 4.5SO 2 / + C 6 H,CH 3 => 7CO lg + 4.5H 2 S° + 4H 2 0 
Toluene oxidation /sulfate reduction 

-142.86 

-597.7 

4.7:1 

0 . 21:1 

5H 2 0 + Ctf HiCH 3 => 2.5C0 2g + 4.5CH 4 

Toluene oxidation / methanogenesis 

-34.08 

-142.6 

0.78:1 

1.28:1 

18 C 2 H 2 CI 4 + C 6 H 5 CH 3 + 14 H 2 0 7 C0 2 + 18 C 2 H 3 CI 3 + 18H + + 18C1' 

Toluene oxidation / PCA reduction 

-382.6 

-1599 

32.8:1 

0.03:1 

I 8 C 2 H 3 CI 3 +C 6 H 5 CH 3 + 14H 2 0=>7C0 2 + I 8 C 2 H 4 CI 2 + 18H + + 18C1‘ 
Toluene oxidation / TCA reduction 

-442.5 

-1850 

26.1:1 

0.04:1 

18 C 2 H 4 CI 2 + C 6 H 5 CH 3 + 14 H 2 0 7 C0 2 + 18 C 2 H 5 CI + 18 H + +18 Cl' 

Toluene oxidation /DCA reduction 

-400.2 

-1673 

19.3:1 

0.05:1 

18C 2 C1 4 + J4H 2 0 + CtHsCHs => 18C 2 HC1 3 + 7CO : + 18 if + 18CI 
Toluene oxidation/ Tetrachloroethylene reductive dehalogenation 

-425.6 

-1779 

32.4:1 

0.03:1 

18C 2 HCl 3 + 14H 2 0 + C6H,CH 3 => 18C 2 H 2 Cl 2 + 7CO : + 18 If + 18CT 
Toluene oxidation/ Trichloroethylene reductive dehalogenation 

-404.9 

-1693 

25.7:1 

0.04:1 

18C 2 H 2 Cl 2 + 14H 2 0 + CtHsCHs => 18C 2 H 3 Cl + 7C0 2 + 18 If + 180 
Toluene oxidation/ cis-Dichloroethylene reductive dehalogenation 

-340.1 

-1422 

18.9:1 

0.05:1 

18C 2 H 3 Cl + 14H 2 0 + CoHsCHs => 18C 2 H 4 + 7C0 2 + 18if + 18 Cl 
Toluene oxidation/ Vinyl chloride reductive dehalogenation 

-331.5 

-1386 

12 .2:1 

0.08:1 

18C 6 Cl 6 + 14H 2 0 + CiHsCHs => 18C(fl,Ch + 7C0 2 + 18lf + 180 
Toluene oxidation/ Hexachlorobenzene reductive dehalogenation 

-410.3 

-1715 

55.6:1 

0 .02:1 

18C6H,Cl 5 + 14H 2 0 + CilfCHi 18C6H 2 Cl 4 + 7C0 2 + 18If 4 180 
Toluene oxidation/ Pentachlorobenzene reductive dehalogenation 

-420.3 

-1757 

48.9:1 

0 .02:1 

18C6H 2 Cl 4 + 14H 2 0 + C6H 5 CH 3 => ISCMCh + 7C0 2 + 18If + 180 
Toluene oxidation/ Tetrachlorobenzene reductive dehalogenation 

-385.2 

-1610 

42.2:1 

0 .02:1 

18C(fl 3 Cl 3 + 14H 2 0 + CiHsCHi => 18C(,H 4 C1 2 + 7CO : + 18 If + 180 
Toluene oxidation/ Trichlorobenzene reductive dehalogenation 

-368.6 

-1541 

35.4:1 

0.03:1 


B3-48 


































Table B.3.7 Continued. 


Coupled Ethylbenzene Oxidation reactions 

AG° r 

kcal/ mole 

AG°r 
kJ/ mole 

Stoichiometric 
Mass Ratio of 
Electron Acceptor 
or Metabolic 
Byproduct to 
Primary Substrate 

Mass of Primary 
Substrate Utilized per 
Mass of Electron 
Acceptor Utilized or 
Metabolic Byproduct 
Produced 

10.502 + C 6 H 5 C:H 5 => 8 CO lg + 5H 2 0 
> Ethylbenzene oxidation /aerobic respiration 

-1066.13 

-4461 

3.17:1 

0.32:1 

8.4NO} + 8.4H* + C 6 H 5 C 2 H 5 => 8 CO:. g + 9.2 H 2 0 + 4.2N 2 . s 
E thylbenzene oxidation /denitrification 

-1080.76 

-4522 

4.92:1 

0.20:1 

46 H* + 22MnO 2 + C 6 H 5 C 2 H 5 => 8 C 02 . g + 22Mn 2+ + 28 H 2 O 

Ethylbenzene oxidation / manganese reduction 

-1066.27 

-4461 

11.39:1 

0.09:1 

84H* + 42Fe(OH) }a + C«}77jC;>/Zj => 8 CO 2 + 42 Fe 2 * + 110H 2 ( 
Ethylbenzene oxidation / iron reduction 

-778.48 

-3257 

22.0:1 

0.05:1 

10.5H* + 5.25SO 2 / + C 6 H 5 C 2 H 5 => 8 CO : . g + 5.25H 2 S 0 + 5H 2 
Ethylbenzene oxidation /sulfate reduction 

-166.75 

-697.7 

4.75:1 

0.21:1 

5.5 H 2 O + C 6 H 5 C 2 H } => 2.75C0 2 . g + 5.25 CH 4 

Ethylbenzene oxidation / methanogenesis , 

-39.83 

-166.7 

0.79:1 

1.27:1 

2 IC 2 H 2 CI 4 + 16H : 0 + C 6 H 5 C 2 H 5 => 21C 2 H } Cl } + 8 CO 2 + 21IF + 2101- 
Ethylbenzene oxidation/ PCA reductive dehalogenation 

-446.43 

-1866 

32.8:1 

0.03:1 

21C 2 H 3 CI 3 + I 6 H 2 O + CiHsCiH, => 2 IC 2 H 4 CI 2 + 8 CO 2 + 2IFF + 2101- 
Ethylbenzene oxidation/ TCA reductive dehalogenation 

-516.36 

-2158 

26.1:1 

0.04:1 

2 IC 2 H 4 CI 2 + 16H 2 0 + C 6 H 5 C 2 H 5 => 2 IC 2 H 5 CI + 8 CO : + 21fF + 2101- 
Ethylbenzene oxidation/DCA reductive dehalogenation 

-467.01 

-1952 

19.4:1 

0.05:1 

2 IC 2 CI 4 + I 6 H 2 O + C 6 H 5 C 2 H 5 => 2 IC 2 HCI 3 + 8 CO 2 + 2 IFF + 21 Or 
Ethylbenzene oxidation/ Tetrachloroethylene reductive dehalogenation 

-496.67 

-2078 

32.8:1 

0.03:1 

2 IC 2 HCI 3 + I 6 H 2 O + CeHsCiHs => 2 IC 2 H 2 CI 2 + 8C0 2 + 211F + 2101- 
Ethylbenzene oxidation/ Trichloroethylene reductive dehalogenation 

-484.70 

-2028 

26.0:1 

0.04:1 

2 IC 2 H 2 CI 2 + I 6 H 2 O + CiHsCiH, => 2 IC 2 H 3 CI + 8 CO 2 + 21fF + 2101- 
Ethylbenzene oxidation/ cis-Dichloroethylene reductive dehalogenation 

-384.74 

-1610 

19.2:1 

0.05:1 

2 IC 2 H 3 CI + I 6 H 2 O + C 6 H 5 C 2 H 5 => 2 IC 2 H 4 + 8 CO 2 + 21IF + 2101- 
Ethylbenzene oxidation/ Vinyl chloride reductive dehalogenation 

-368.79 

-1617 

12.3:1 

0.08:1 

21C 6 Cl 6 + I 6 H 2 O + C 6 H 5 C 2 H 5 => 2lC^Ch + 8C0 2 + 21fF + 21CI 
Ethylbenzene oxidation/ Hexachlorobenzene reductive dehalogenation 

-478.7 

-2001 

55.6:1 

0.02:1 

21 C 6 HiCl s + I 6 H 2 O + C 6 H s C 2 H 5 => 21C 6 H 2 Cl4 + 80O 2 + 211F + 2101- 
Ethylbenzene oxidation/ Pentachlorobenzene reductive dehalogenation 

-490.4 

-2050 

48.9:1 

0.02:1 

2 IC 6 H 2 CI 4 + I 6 H 2 O + C 6 H 5 C 2 H 5 => 2 IC 6 H 3 CI 3 + 8 CO 2 + 21FF + 21CI 

Ethylbenzene oxidation/ Tetrachlorobenzene reductive dehalogenation 

-449.4 

-1878 

42.2:1 

0.02:1 

21C6H3C13 + I 6 H 2 O + CeHsCiHs => 27QZZ<C/.> + 8C0 2 + 21lF + 2101- 

Ethylbenzene oxidation/ Trichlorobenzene reductive dehalogenation 

-430.1 

-1794 

35.5:1 

0.03:1 


B3-49 









































Table B.3. 7 Continued. 


Coupled m-Xylene Oxidation Reactions 

AG° r 

(kcal/ mole) 

AG° r 

(kJ/ mole) 

Stoichiometric Mass 
Ratio of Electron 
Acceptor or 
Metabolic 
Byproduct to 
Primary Substrate 

Mass of Primary 
Substrate Utilized per 
Mass of Electron 
Acceptor Utilized or 
Metabolic Byproduct 
Produced 

10.5 02 +C 6 H 4 (CH 3 )} => 8C0 2 + 5 H 2 0 
m-Xylene oxidation / aerobic respiration 

-1063.25 

-4448 

3.17:1 

0.32:1 

8.4 H*+ 8.4NO' } + C 6 H 4 (CH 3 ) 2 => 8C0 2 + 4.2 N 2 + 9.2 H 2 0 
m-Xylene oxidation /denitrification 

-1077.81 

-4509 

4.92:1 

0.20:1 

46 H*+ 22Mn0 2 + C 6 H 4 (CH 3 ) 2 => 8C0 2 + 22 Mn 2 * + 28 H 2 0 

m-Xylene oxidation / manganese reduction 

-1063.39 

-4449 

11.39:1 

0.09:1 

84 H*+ 42Fe(OH) } , a + C 6 H 4 (CH 3 ) 2 => 8C0 2 + 42 Fe 2 * + 110 H 2 0 

m-Xylene oxidation / iron reduction 

-775.61 

-3245 

22:1 

0.05:1 

10.5H* + 5.25S0 4 2 ' + C 6 H 4 (CH}) 2 => 8C0 2 + 5.25 H 2 S°+ 5 H 2 0 

m-Xylene oxidation /sulfate reduction 

-163.87 

-685.6 

4.75:1 

0.21:1 

5.5H 2 0 + C 6 H 4 (CH 3 ) 2 => 2.75CO } + 5.25CH 4 
m^Xylene oxidation /methanogenesis 

-36.95 

-154.6 

0.79:1 ^ 

1.27:1 

21C 2 H 2 Cl 4 + 16H 2 0 + C«H 4 (CH 3 ) 2 => 21C 2 H 3 Cl 3 + 8C0 2 + 21H* + 

21 Cl' 

m-Xylene oxidation/ PCA reductive dehalogenation 

-445.70 

-1863 

32.7:1 

0.03:1 

21C 2 H 3 Cl 3 + 16H 2 0 + C 6 H 4 (CH 3 ) 2 => 21C 2 H 4 Cl 2 + 8CO : + 21H* + 

21 Cl' 

m-Xylene oxidation / TCA reductive dehalogenation 

-513.48 

-2146 

26.0:1 

0.04:1 

21 C 2 H 4 Cl 2 + 16H 2 0 + CiH 4 (CH 3 ) 2 => 21C 2 H 5 Cl + 8C0 2 + 21H* + 

2 ICY 

m-Xylene oxidation/ DCA reductive dehalogenation 

-464.13 

-1940 

19.3: 

0.05:1 

21C 2 Cl 4 + 16H 2 0 + C <s H 4 (CH j ) 2 => 21C 2 HCl 3 + 8C0 2 + 21H* + 21CY 
m-Xylene oxidation/ Tetrachloroethylene reductive dehalogenation 

-493.79 

-2066 

32.8:1 

0.03:1 

21C 2 HCl 3 + 16H 2 0 + C6H 4 (CH 3 ) 2 => 21C 2 H 2 Cl 2 + 8C0 2 + 21H* + 21CY 
m-Xylene oxidation/ Trichloroethylene reductive dehalogenation 

-469.59 

-1963 

26.0:1 

0.04:1 

21C 2 H 2 C1 2 + 16H 2 0 + C6H 4 (CH 3 ) 2 => 21C 2 H 3 Cl + 8C0 2 + 21H* + 

2 ICY 

m-Xylene oxidation/ cis-Dichloroethylene reductive dehalogenation 

-393.99 

-1647 

19.2:1 

0.05:1 

21C 2 H 3 Cl + 16H 2 0 + CtH 4 (CH 3 ) 2 => 21C } H 4 + 8C0 2 + 21H* + 21CY 
m-Xylene oxidation/ Vinyl chloride reductive dehalogenation 

-383.91 

-1605 

12.3:1 

0.08:1 

21C 6 Cl 6 + 16H 2 0 + CiH 4 (CH 3 ) 2 => 21C6H,Cl 5 + 8C0 2 + 21H * + 21CI 
m-Xylene oxidation/Hexachlorobenzene reductive dehalogenation 

-475.9 

-1989 

55.6:1 

0.02:1 

21C 6 H,CI s + 16H 2 0 + ChH 4 (CH 3 ) 2 => 21C6H 2 Cl 4 + 8C0 2 + 21H * + 21CI 
m-Xylene oxidation/ Pentachlorobenzene reductive dehalogenation 

-487.5 

-2038 

48.9:1 

0.02:1 

21C 6 H 2 Cl 4 + 16H 2 0 + C6H 4 (CH 3 ) 2 => 270,,//^ + 8C0 2 + 21H' + 21CI 
m-Xylene oxidation/ Tetrachlorobenzene reductive dehalogenation 

-446.6 

-1867 

42.2:1 

0.02:1 

21CsH 3 Cl 3 + 16H 2 0 + CJtY 4 (CH 3 ) 2 => 21C 6 H 4 Cl 2 + 8CO : + 21H * + 

21CF 

m-Xylene oxidation/ Trichlorobenzene reductive dehalogenation 

-426.9 

-1784 

35.5:1 

0.03:1 


B3-50 





























Table B.3. 7 Continued. 


Coupled Naphthalene Oxidation Reactions 

AG°, 

(kcal/ 

mole) 

AG\ 

(kJ/ 

mole) 

Stoichiometric Mass 
Ratio of Electron 
Acceptor or Metabolic 
Byproduct to Primary 
Substrate 

M ass of Primary 
Substrate Utilized per 
Mass of Electron 
Acceptor Utilized or 
Metabolic Byproduct 
Produced 

120 2 + C I0 H S => 10CO 2 + 4H 2 0 

Naphthalene oxidation /aerobic respiration 

-1217.40 

-5094 

3.00:1 

0.33:1 

* 9.6NO$~ + 9.6H* + CioHs => IOCO 2 + 8 . 8 H 2 O + 4 . 8 N 2 . 

Naphthalene oxidation / denitrification 

-1234.04 

-5163 

4.65:1 

0.22:1 

24Mn0 2 + 48H * + CioHs => 10CO 2 + 24Mn 2 * + 28H 2 0 

Naphthalene oxidation /manganese reduction 

-1217.57 

-5094 

16.31:1 

0.06:1 

48Fe(OH) } . a + 96H* + C I0 H S => 10CO 2 + 48Fe 2 * + 124H 2 0 
aphthalene oxidation / iron reduction 

-932.64 

-3902 

40.13:1 

0.02:1 

6 SO 4 2 + 12H * + CioHs 10CO 2 + 6 H 2 S° + 4H 2 0 

Error! Switch argum ettt not specified.Naphthalene oxidation / 
sulfate reduction 

-196.98 

-824.2 

4.50:1 

0.22:1 

8 H 2 0 + CioHs => 4C0 2 + 6 CH 4 

Nap hthalene oxidation / methanogenesis 

-44.49 

-186.1 

1.13:1 

0.88:1 

24C 2 H 2 CI 4 + 20H : O + CioHs => 24C 2 H 3 Cl } + 10CO 2 + 24H* + 

24CC 

N aphthalene oxidation/ PC A reductive dehalogenation 

-511.68 

* 

-2139 

31.1:1 

0.03:1 

24C 2 H}Cl } + 20H 2 O + CioHs => 24C 2 H,Cl 2 + 10CO 2 + 24K* + 

24CT 

Naphthalene oxidation/ TCA reductive dehalogenation 

-589.09 

-2462 

24.8:1 

0.04:1 

24C 2 H 4 Cl 2 + 20H 2 O + CioHs => 24C 2 H s Cl + 10CO 2 + 24H* + 

24CC 

N aphthalene oxidation/ DCA reductive dehalogenation 

-532.69 

-2227 

18.4:1 

0.05:1 

24C 2 Cl 4 + 20H 2 O + CioHs => 24C 2 HCl 3 + 10CO 2 + 24H + + 24C/' 

Naphthalene oxidation/ Tetrachloroethylene reductive 
dehalogenation 

-566.59 

-2371 

31.1:1 

0.03:1 

24C 2 HCl 3 + 20H 2 O + => 24C 2 H 2 Cl 2 + 10CO 2 + 24//* + 24CV 

Naphthalene oxidation/ Trichloroethylene reductive dehalogenation 

-552.91 

-2313 

24.6:1 

0.04:1 

24C 2 H 2 Cl 2 + 20H 2 O + C, 0 Hs => 24C 2 H 3 Cl + /0CO, + 24H* + 24CT 

Naphthalene oxidation/cis-Dichloroethylene reductive 
dehalogenation 

-438.67 

-1835 

18.2:1 

0.05:1 

24C 2 H 3 Cl + 2 0H 2 O + CioHs => 24C 2 H t + 10CO 2 + 24H * +24CT 

Nap hthalene oxidation/ Vinyl chloride reductive dehalogenation 

-441.01 

-1843 

11.6:1 

0.09:1 

24C 6 CI 6 + 20H 2 O + C, 0 Hs => 24C 6 HiCl s + 10CO 2 + 24H'+24Cl 

Naphthalene oxidation/ Hexachlorobenzene reductive 
dehalogenation 

-545.94 

-2282 

52.9:1 

0.02:1 

24C sH iCI 3 + 20H 2 O + C,oH s => 24C tH 2 Cl 10CO 2 + 24H* + 24CI 

Naphthalene oxidation/Pentachlorobenzene reductive 
dehalogenation 

-559.33 

-2338 

46.5:1 

0.02:1 

24CsH 2 Cl 4 + 20H 2 O + C, 0 Hs => 24C 6 H 3 Cl 3 + 10CO 2 + 24H * + 

24CI 

Naphthalene oxidation/ Tetrachlorobenzene reductive 
dehalogenation 

-512.53 

-2142 

40.1:1 

0.02:1 

24C6H 3 Cl 3 + 2 OH 2 0 + C;o//« => 24C s H 4 Cl 2 + 10CO 2 + 24H* + 

24CF 

Naphthalene oxidation/ Trichlorobenzene reductive dehalogenation 

-490.45 

-2050 

33.8:1 

0.03:1 


B3-51 































Table B. 3.7 Continued. 


Coupled 1,3,5-Trimethylbenzene (1,3,5-TMB) Oxidation Reactions 

AG° r 

(kcal/ mole) 

AG° r 

(kJ/ mole) 

Stoichiometric Mass 
Ratio of Electron 
Acceptor or Metabolic 
Byproduct to Primary 
Substrate 

Mass of Primary Substrate 
Utilized per Mass of 
Electron Acceptor Utilized 
or Metabolic Byproduct 
Produced 

120 2 + C6H 3 (CH 3 )s => 9C0 2 + 6 H 2 0 

1,3,5-TMB oxidation /aerobic respiration 

-1213.29 

-5076 

3.20:1 

0.31:1 

9.6N0 3 + 9.6FF + C6H 3 (CH 3 ) 3 => 9C0 2 + 10.8H 2 O + 4.8N 2 , g 

1,3,5-TMB oxidation / denitrification 

-1229.93 

-5146 

4.96:1 

0.20:1 

24MnO : + 481T + C6H 3 (CH 3 ) 3 => 9C0 2 + 30H 2 O + 24Mn :+ 

1,3,5-TMB oxidation / manganese reduction 

-1213.46 

-5077 

17.40:1 

0.06:1 

48Fe(OH) 3 , a + 96tf + C6H 3 (CH 3 ) 3 => 9C0 2 + 48Fe 2 * + 126H 2 0 

1,3,5-TMB oxidation /iron reduction 

-928.53 

-3885 

42.80:1 

0.02:1 

6S0 4 2 ‘ + 12 H* + C6H 3 (CH 3 ) 3 => 9C0 2 + 6H 2 0 + 6 H 2 S° 

1,3,5-TMB oxidation /sulfate reduction 

-192.87 

-807.0 

4.80:1 

0.21:1 

6H 2 0 + C 6 H 3 (CH 3 ) 3 => 3C0 2 + <5C//, 

1,3,5-TMB oxidation / methanogenesis 

-40.39 

-169.0 

0.90:1 

1.11:1 

24 C 2 H 2 Cl 4 + 18H 2 0 + C6H 3 (CH 3 ) 3 => 24C 2 H 3 C1 3 + 9C0 2 + 24H* + 24CI 
1,3,5-TMB oxidation/PCA reductive dehalogenation 

-507.36 

-2121 

33.2:1 

0.03:1 

24C 2 H 3 C1 3 + 18H 2 0 + C6H 3 (CH 3 ) 3 => 24C 2 H 4 Cl 2 + PC0 2 + 24FT + 24CT 
1,3,5-TMB oxidation/ TCA reductive dehalogenation 

-584.99 

-2445 

26.4:1 

0.04:1 

24C : H 4 Cl 2 + 18H 2 0 + C(JH 3 (CH 3 ) } => 24C 2 H 5 C1 + 9CO : + 24FT + 24CT 
1,3,5-TMB oxidation/ DCA reductive dehalogenation 

-528.59 

-2210 

19.6:1 

0.05:1 

24C 2 C1 4 + ;S// 2 0 + C6H 3 (CH 3 ) } => 24C 2 HC1 3 + 9C0 2 + 24FF + 

1,3,5-TMB oxidation/ Tetrachloroethene reductive dehalogenation 

-562.48 

-2353 

33.2:1 

0.03:1 

24C 2 HCl } + 18H 2 0 + CtHifCH}) 3 => 24C 2 H 2 Cl 2 + PC<9 2 + 24IT+ 24Ct 
1,3,5-TMB oxidation/ Trichloroethene reductive dehalogenation 

-548.80 

-2296 

26.3:1 

0.04:1 

24C 2 H 2 Cl 2 + 18H 2 0 + C6H 3 (CH 3 ) 3 => 24C 2 H 3 Cl + 9C0 2 + 241F + 24CT 
1,3,5-TMB oxidation/cis-Dichloroethene reductive dehalogenation 

-434.56 

-1818 

19.4:1 

0.05:1 

24C 2 H 3 Cl + 18H 2 0 + C 6 H 3 (CH 3 ) 3 => 24C 2 H 4 + 9C0 2 + 24FF + 24CT 
1,3,5-TMB oxidation/ Vinyl chloride reductive dehalogenation 

-436.91 

-1826 

12.4:1 

0.08:1 

24C 6 Cl 6 + 18H 2 0 + C6H 3 (CH 3 ) 3 => 24C6H,Cl, + 9C0 2 + 24fF + 24CT 
1,3,5-TMB oxidation/Hexachlorobenzene reductive dehalogenation 

-541.84 

-2265 

56.4:1 

0.02:1 

} 

24C 6 H 1 Cl 5 + 18H 2 0 + C 6 H 3 (CH 3 ) 3 => 24C6H 2 Cl 4 + 9C0 2 + 24FT + 24CT 
1,3,5-TMB oxidation/Pentachlorobenzene reductive dehalogenation 

-555.23 

-2321 

49.6:1 

0.02:1 

24C(jFl 2 Cl 4 + 18H 2 0 + C6H 3 (CH 3 ) 3 => 24C6H 3 Cl 3 + 9C0 2 + 24FF + 24CT 

1,3,5-TMB oxidation/ Tetrachlorobenzene reductive dehalogenation 

-508.43 

-2125 

42.8:1 

0.02:1 

24CaH 3 Cl 3 + 18H 2 0 + C6H 3 (CH 3 ) 3 => 24C6H 4 Cl 2 + 9C0 2 + 24H* + 24CT 

1,3,5-TMB oxidation/ Trichlorobenzene reductive dehalogenation 

-486.35 

-2033 

36.0:1 

0.03:1 


B3-52 































Table B. 3.7 Continued. 


Coupled 1,2,4-Trimethylbenzene (1,2,4-TMB) Oxidation Reactions 

AG° r 

(kcal/ mole) 

AG° r 

(kJ/ mole) 

Stoichiometric Mass 
Ratio of Electron 
Acceptor or 
Metabolic 
Byproduct to 
Primary Substrate 

Mass of Primary Substrate 
Utilized per Mass of 
Electron Acceptor Utilized 
or Metabolic Byproduct 
Produced 

120 2 + C 6 H)(CHi)i => 9C0 2 + 6H 2 0 

1,2,4-TMB oxidation /aerobic respiration 

-1212.92 

-5075 

3.20:1 

0.31:1 

9.6NO) + 9.6H+ + CsHsfCHi)} => 9C0 2 + 10.8H 2 O + 4.8N 2 . g 

1,2,4-TMB oxidation / denitrification 

-1229.56 

-5144 

4.96:1 

0.20:1 

24MnO ? + 48H + + CeHsfCHih => 9C0 2 + 30H 2 O + 24Mn 2+ 

1,2,4-TMB oxidation / manganese reduction 

-1213.09 

-5076 

17.4:1 

0.06:1 

48Fe(OH)t „ + 96H* + CJU/CHiU => 9CO, + 48Fe 2 * + 126H,0 

1,2,4-TMB oxidation / iron reduction 

-928.16 

-3883 

42.8:1 

0.02:1 

6S0 4 2 ' + 12 Tt + C 6 H}(CH } ) } => 9C0 2 + 6H 2 0 + 6 H 2 S° 

1,2,4-TMB oxidation /sulfate reduction 

-192.50 

-805.4 

4.80:1 

0.21:1 

6H 2 0 + CsHifCHih => 3C0 2 + 6 CH 4 

1,2,4-TMB oxidation / methanogenesis > 

-40.02 

-167.4 

0.90:1 

1.11:1 

24C 2 H 2 CU + 18H 2 0 + CcHfCHj)} => 24C 2 H 3 Cl 3 + 9C0 2 + 24fF + 

24CT 

1,2,4-TMB oxidation/ PCA reductive dehalogenation 

-507.36 

-2121 

33.2:1 

0.03:1 

24C 2 H 3 Cl 3 + 18H 2 0 + C6H 3 (CH 3 ) 3 => 24C 2 H 4 C1 2 + 9C0 2 + 24H* + 

24 cr 

1,2,4-TMB oxidation/ TCA reductive dehalogenation 

-584.62 

-2444 

26.4:1 

0.04:1 

24C 2 H 4 Cl 2 + 18H 2 0 + C6H 3 (CH 3 ) 3 => 24C 2 H 3 Cl + 9C0 2 + 24H + + 24CT 
1,2,4-TMB oxidation/DCA reductive dehalogenation 

-528.22 

-2208 

19.6:1 

0.05:1 

24C 2 Cl 4 + 18H 2 0 + C6H 3 (CH 3 ) } => 24C 2 HC1 3 + 9C0 2 + 24H+ + 24CT 
1,2,4-TMB oxidation/ PCE reductive dehalogenation 

-562.11 

-2352 

33.2:1 

0.03:1 

24C 2 HCl 3 + 18H 2 0 + C6H 3 (CH 3 ) 3 => 24C 2 H 2 Cl 2 + 9CO : + 24FT + 24CV 
1,2,4-TMB oxidation/ TCE reductive dehalogenation 

-548.43 

-2295 

26.3:1 

0.04:1 

24C 2 H 2 Cl 2 + 18H 2 0 + CiH 3 (CH 3 ) 3 => 24C 2 H 3 Cl + 9C0 2 + 24H+ + 24CY 
1,2,4-TMB oxidation/ cis-DCE reductive dehalogenation 

-434.19 

-1817 

19.4:1 

0.05:1 

24C 2 H 3 Cl + 18H 2 0 + C6H 3 (CH 3 ) 3 => 24C 2 H 4 + 9C0 2 + 24H* + 24CT 

1,2,4-TMB oxidation/ Vinyl chloride reductive dehalogenation 

-436.54 

-1825 

12.4:1 

0.08:1 

24C 6 Cl 6 + 18H 2 0 + C6H 3 (CH 3 ) 3 => 24C6H,C1 5 + 9C0 2 + 24H* + 24CT 
1,2,4-TMB oxidation/Hexachlorobenzene reductive dehalogenation 

-541.47 

-2263 

56.4:1 

0.02:1 

24C6H,Cl 5 + 18H 2 0 + C 6 H 3 (CH 3 ) 3 => 24C 6 H 2 Cl 4 + 9C0 2 + 241T + 24CT 
1,2,4-TMB oxidation/ Pentachlorobenzene reductive dehalogenation 

-554.86 

-2319 

49.6:1 

0.02:1 

24CsH 2 Cl 4 + 18H 2 0 + C6H 3 (CH 3 ) 3 => 24CaH 3 Cl 3 + 9CO : + 24H* + 24CT 

1,2,4-TMB oxidation/ Tetrachlorobenzene reductive dehalogenation 

-508.06 

-2124 

42.8:1 

0.02:1 

24CsH 3 CI 3 + 18H 2 0 + C6H 3 (CH 3 ) 3 => 24C,sH 4 Cl 2 + 9CO : + 24H+ + 24Ct 

1,2,4-TMB oxidation/ Trichlorobenzene reductive dehalogenation 

-485.98 

-2031 

36.0:1 

0.03:1 


B3-53 

































Table B. 3.7 Continued. 


Coupled Vinyl Chloride Oxidation Reactions 

AG° r 

(kcal/ mole) 

AG°r 

(kJ/ mole) 

Stoichiometric Mass 
Ratio of Electron 
Acceptor or Metabolic 
Byproduct to Primary 
Substrate 

Mass of Primaiy Substrate 
Utilized per Mass of 
Electron Acceptor Utilized 
or Metabolic Byproduct 
Produced 

2.50 2 + C 2 HiCl => 2CO 2 + H 2 0 + ft + Cl 

Vinyl Chloride oxidation / aerobic respiration 

-288.98 

-1209 

1.29:1 

0.78:1 

2NOy + ft C 2 H 3 CI => 2C0 2 + 2H 2 0 + Ct + N 2 . g 

Vinyl Chloride oxidation / denitrification 

-292.44 

-1224 

2.00:1 

0.50:1 

5Mn0 2 + 9ft + C 2 H } Cl => 2C0 2 + 6H 2 0 + 5Mn 2+ + Cl 

Vinyl Chloride oxidation / manganese reduction 

-289.01 

-1209 

7.02:1 

0.14:1 

lOFefOH), . + 19ft + CJhfCHAi => 2CCh + lOFe 2 * + 26H>0 + Cl 
Vinyl Chloride oxidation / iron reduction 

-229.65 

-960.9 

17.3:1 

0.06:1 

1.25S0 4 2 ' + 1.5H + + C 2 H 3 CI => 2C0 2 + H 2 0 + 1.25H 2 S° + CT 

Vinyl Chloride oxidation /sulfate reduction 

-76.40 

-319.7 

1.94:1 

0.52:1 

1.5 H 2 0 + C 2 H 3 CI => .75C0 2 + 1.25CH 4 + ft + Cl 

Vinyl Chloride oxidation / methanogenesis 

-44.62 

-186.7 

0.44:1 

2.27:1 

5C 2 H 2 Cl 4 + 4H 2 0 + C 2 H 3 CI => 5 C 2 H 3 CI 3 + 2CO : + 6 ft + 6 Ct 

Vinyl Chloride oxidation/ PCA reductive dehalogenation 

-141.90 

-593.1 

13.4:1 

0.07:1 

5 C 2 H 3 CI 3 + 4H 2 0 + C 2 ff } Cl => 5C 2 H 4 Cl 2 + 2C0 2 + 6 ft + 6 Ct 

Vinyl Chloride oxidation/ TCA reductive dehalogenation 

-158.08 

-661 

10.7:1 

0.09:1 

5C 2 H 4 Cl 2 + 4H : 0 + C 2 H 3 Cl => 5C 2 H 5 Cl + 2C0 2 + 6 ft + 6 Ct 

Vinyl Chloride oxidation/ DCA reductive dehalogenation 

-146.33 

-612 

7.92:1 

0.13:1 

5C 2 C1 4 + 4H 2 0 + C 2 H 3 CI => 5C 2 HCl 3 + 2C0 2 + 6 ft + 6 CI 

Vinyl Chloride oxidation/DCE reductive dehalogenation 

-153.39 

-641.8 

13.4:1 

0.07:1 

5C 2 HCl 3 + 4H 2 0 + C 2 H 3 CI => 5C 2 H 2 Cl 2 + 2C0 2 + 6 ft + 6 Ct 

Vinyl Chloride oxidation/ TCE reductive dehalogenation 

-150.54 

-629.9 

10.6:1 

0.09:1 

5C 2 H 2 Cl 2 + 4H 2 0 + C 2 H 3 CI => 5C 2 HsCl + 2C0 2 + 6 ft + 6 Ct 

Vinyl Chloride oxidation/ cis-DCE reductive dehalogenation 

-126.74 

-530.3 

7.82:1 

0.13:1 

5C 6 Cl 6 + 4H 2 0 + C 2 H 3 CI => SCfFfiCh + 2C0 2 + 6 ft + 6 Ct 

Vinyl Chloride oxidation/ Hexachlorobenzene reductive dehalogenation 

-144.60 

-604.4 

22.8:1 

0.04:1 

SCffhCls + 4H 2 0 + C 2 H 3 CI => 5C6H 2 Cl 4 + 2CO : + 6 ft + 6 CI 

Vinyl Chloride oxidation/Pentachlorobenzene reductive dehalogenation 

-138.59 

-579.3 

20.0:1 

0.05:1 

5C 6 H 2 Cl 4 + 4H 2 0 + C 2 H 3 CI => SCfifiCf + 2C0 2 + 6 ft + 6 Ci 

Vinyl Chloride oxidation/ Tetrachlorobenzene reductive dehalogenation 

-142.13 

-594.1 

17.3:1 

0.06:1 

5C 6 H 3 Cl 3 + 4H 2 0 + C 2 H 3 CI => 5C 6 H 4 Cl 2 + 2C0 2 + 6 ft + 6 CI' 

Vinyl Chloride oxidation/ Trichlorobenzene reductive dehalogenation 

-137.53 

-574.9 

14.5:1 

0.07:1 

20 2 + C 2 H 2 Cl 2 => 2C0 2 + 2ft + 2Ct 

DCE oxidation /aerobic^es^imtior^^ 

-256.53 

-1072 

1.31:1 

0.76:1 


B3-54 
































Table B. 3.7 Continued. 


Coupled Chlorobenzene Oxidation Reactions 

AG°r 

(kcal/ 

mole) 

AG°r 

(kJ / mole) 

Stoichiometric 
Mass Ratio of 
Electron 
Acceptor or 
Metabolic 
Byproduct to 
Primary 
Substrate 

Mass of Primary 
Substrate Utilized 
per Mass of 
Electron Acceptor 
Utilized or 
Metabolic 
Byproduct 
Produced 

70: + C6H 4 Cl => 6C0 2 + /T + 2H 2 0+ Cl 

Chlorobenzene oxidation /aerobic respiration 

-731.62 

-3061 

2.00:1 

0.50:1 

5 . 6 NO 3 - + 4.6fF + CtHiCl => 6C0 2 + 4 . 8 H 2 O + 2.8N 2 , g + 2Ct 
Chlorobenzene oxidation / denitrification 

-741.33 

-3102 

3.10:1 

0.32:1 

l4Mn0 2 + 27tF + CMsCl => 6CO : + 16H 2 0 + 14Mn 2+ + Ct 
Chlorobenzene oxidation / manganese reduction 

-731.72 

-3062 

10.9:1 

0.09:1 

28Fe(OH), „ + 55FF + CM,Cl => 6CO, + 72H,0 + 28Fe 2+ + Ct 

Chlorobenzene oxidation / iron reduction 

-565.51 

-2366 

26.8:1 

0.04:1 

3.5S0 2 ' + 6tF + CMsCl => 6C0 2 + 2H 2 0 + 3.5H^ + Ct 
Chlorobenzene oxidation / sulfate reduction 

-136.38 

-570.6 

3.00:1 

0.33:1 

5H : 0 + CMsCl => 2.SCO 2 + 3.5CH 4 + H* + Ct 
Chlorobenzene oxidation / methanogenesis 

-47.43 

-198.4 

0.80:1 

1.25:1 

14C 2 H 2 Cl 4 + 12H : 0 + C6HsCl=> 14C 2 H 3 Cl 3 + 6C0 2 + 15 IF + 15Ct 
Chlorobenzene oxidation/ PC A reductive dehalogenation 

-320.04 

-1338 

20.8:1 

0.05:1 

14C 2 H 3 Cl 3 + 12H 2 0 + C6H 5 C1=> 14C 2 H 4 Cl 2 + 6C0 2 + 15H* + 15Ct 
Chlorobenzene oxidation/ TCA reductive dehalogenation 

-365.11 

-1526 

16.5:1 

0.06:1 

14C 2 H 4 Cl 2 + 12H 2 0 + C 6 HsCl => 14C 2 H 5 Cl + 6C0 2 + 15IF + 15CI 
Chlorobenzene oxidation/ DC A reductive dehalogenation 

-332.21 

-1389 

12.3:1 

0.08:1 

14C 2 CI 4 + 12H 2 0 + CMsCl => 14C 2 HCl 3 + 6C0 2 + 15 Ft + 15Ct 
Chlorobenzene oxidation/ PCE reductive dehalogenation 

-351.99 

-1473 

20.7:1 

0.05:1 

14C 2 HCl 3 + 12H 2 0 + CMsCl => 14C 2 H 2 Cl2 + 6C0 2 + 15FT + 15Ct 
Chlorobenzene oxidation/ TCE reductive dehalogenation 

-344.01 

-1439 

16.4:1 

0.06:1 

14C 2 H 2 Cl2 + 12H 2 0 + CMsCl => 14C 2 H 3 Cl + 6C0 2 + 15H* + 15CI 
Chlorobenzene oxidation/ cis-DCE reductive dehalogenation 

-277.37 

-1161 

12.1:1 

0.08:1 

14C 2 H 3 Cl + 12H 2 0 + CMsCl => 14C 2 H 4 + 6C0 2 + 15tF + 15CI 

Chlorobenzene oxidation/ Vinyl chloride reductive dehalogenation 

-278.73 

-1165 

7.75:1 

0.13:1 


B3-55 































Table B. 3 .7 Continued. 


Coupled Dichlorobenzene Oxidation Reactions 

AG°, 

(kcal/ mole) 

AG° r 

(kJ/ mole) 

Stoichiometric 
Mass Ratio of 
Electron 
Acceptor or 
Metabolic 
Byproduct to 
Primary 
Substrate 

Mass of Primary 
Substrate Utilized 
per Mass of 
Electron Acceptor 
Utilized or 
Metabolic 
Byproduct 
Produced 

6.50 2 + C 6 H 4 CI 2 => 6C0 2 + 2Ff + H 2 0+ 2C1- 
Dichlorobenzene oxidation /aerobic respiration 

-698.36 

-2919 

1.42:1 

0.70:1 

5.2NO j + 3.2/T + CeHiCh => 6 CO : + 3.6H 2 0 + 2.6N 2 , g + 2Ci 
Dichlorobenzene oxidation / denitrification 

-708.76 

-2963 

1.64:1 

0.61:1 

13Mn0 2 + 24tT + CJijCh => 6C0 2 + 14H 2 0 + 13Mn 2 * + 2Ct 
Dichlorobenzene oxidation /manganese reduction 

-698.36 

-2919 

7.75:1 

0.13:1 

26Fe(OH)\ n + SOFT + CM 4 CI 1 => dCO; + 66 H 2 O + 26fe 2+ + 2CI 

Dichlorobenzene oxidation / iron reduction 

-521.56 

-2180 

19.05:1 

0.05:1 

3.25SO/- + 4.5Pf + CMCl^ 6CO : + H 2 0 + 3 . 25 H 2 S 0 + 2Ct 
Dichlorobenzene oxidation /sulfate reduction 

-142.74 

-596.7 

2.14:1 

0.47:1 

5.5H 2 0 + C 6 H 4 CI 2 => 2.75 CO 2 + 3.25CH 4 + 2/T + 2Ct 

Dichlorobenzene oxidation / methanogenesis 

-64.22 

-268.4 

0.33:1 

2.99:1 

I 3 C 2 H 2 CI 4 + 12H 2 0 + C6H4Cl 2 => 13C 2 H } Cl} + + 15 IT + 15Ct 

Dichlorobenzene oxidation/ PCA reductive dehalogenation 

-317.20 

-1326 

14.8:1 

0.07:1 

13C 2 H}Cli + 12H 2 0 + C 6 H 4 Cl 2 => 13C 2 H 4 C1 2 + 6C0 2 + 15 If + 15Ct 
Dichlorobenzene oxidation/ TCA reductive dehalogenation 

-358.93 

-1500 

11.8:1 

0.09:1 

I 3 C 2 H 4 CI 2 + 12H : 0 + C6H 4 Cl2 => 13C 2 H 5 Cl + 6C0 2 + 15 If + 15 Cl 
Dichlorobenzene oxidation/ DCA reductive dehalogenation 

-328.38 

-1373 

8.73:1 

0.11:1 

I 3 C 2 CI 4 + 12H 2 0 + C 6 H 4 Cl 2 => 13C 2 HCl 3 + dC0 2 + 151f + 75C/ 
Dichlorobenzene oxidation/ PCE reductive dehalogenation 

-347.10 

-1450 

14.6:1 

0.07:1 

13C 2 HCl 3 + 12H 2 0 + C 6 H 4 Cl 2 => I 3 C 2 H 2 CI 2 + 6 CO 2 + 15If + 15Cf 
Dichlorobenzene oxidation/ TCE reductive dehalogenation 

-339.56 

-1419 

11.6:1 

0.09:1 

13C 2 H 2 Cl2 + 12H : 0 + C 6 H 4 Cl 2 => 13C 2 H 3 CI + 6C0 2 + 15 If + 15Ct 
Dichlorobenzene oxidation/ cis-DCE reductive dehalogenation 

-277.68 

-1161 

8.55:1 

0.12:1 

13C 2 H } C1 + 12H : 0 + C(fltCf => BC 2 H 4 + 6 CO 2 + 15If + 15CI 
Dichlorobenzene oxidation/ Vinyl chloride reductive dehalogenation 

-278.72 

-1165 

5.52:1 

0.18:1 


B3-56 
























Table B. 3.7 Continued. 


Coupled Trichlorobenzene Oxidation Reactions 

AG°, 

(kcal/ mole) 

AG° r 

(kJ / mole) 

Stoichiometric Mass 
Ratio of Electron 
Acceptor or Metabolic 
Byproduct to Primary 
Substrate 

Mass of Primary 
Substrate Utilized per 
Mass of Electron 
Acceptor Utilized or 
Metabolic Byproduct 
Produced 

6 O 2 + CeHsCh => 6 CO: + 3FF + 3Ct 

Trichlorobenzene oxidation /aerobic respiration 

-668.16 

-2793 

1.07:1 

0.94:1 

4.8NO) + 1.8FT + CeHsCU => 6C0 2 + 2.4H 2 0 + 2.4N 2 . g + 3Ct 
Trichlorobenzene oxidation / denitrification 

-677.76 

-2833 

1.65:1 

0.60:1 

12Mn0 2 + 2llT + CfUsCli- => 6C0 2 + 12H 2 0 + J2Mn 2+ + 3CI 
Trichlorobenzene oxidation / manganese reduction 

-688.16 

-2793 

5.80:1 

0.17:1 

24Fe/OH)i* + 45FT + C<H*Cl/=z 6 CO , + 60H,O + 24Fe 2+ + 3Ct 

Trichlorobenzene oxidation / iron reduction 

-504.96 

-2111 

14.3:1 

0.07:1 

3SO/ + 3Fr + C<//iC/i=> dCO; + + 3CV 

Trichlorobenzene oxidation / sulfate reduction 

-155.28 

-649.1 

1.60:1 

0.63:1 

6H 2 0 + CsHiCh => 3C0 2 + 3CH 4 + 3fF + 3CT 
Trichlorobenzene oxidation / methanogenesis 

-82.80 

> 

-346.1 

0.25:1 

4.00:1 

12C 2 H 2 Cl 4 + 12 H 2 0 + C 6 H}CU=> 12C 2 H 3 Cl 3 + 6C0 2 + 75/T t 15Cl 
Trichlorobenzene oxidation/ PCA reductive dehalogenation 

-316.32 

-1322 

11.1:1 

0.09:1 

12C 2 H 3 Cl 3 + 12H 2 0 + C 6 H 3 Cl 3 => 12C 2 H 4 Cl 2 + 6C0 2 + 15fF + 15CT 
Trichlorobenzene oxidation/ TCA reductive dehalogenation 

-354.82 

-1483 

8.8:1 

0.11:1 

12 C 2 H 4 C 1 2 + 12 H 2 0 + C(JH 3 Cl } => 12 C 2 H 5 Cl + 6 CO : + 15 FT + i5cr 
Trichlorobenzene oxidation/DCA reductive dehalogenation 

-326.62 

-1365 

6.53:1 

0.15:1 

12C 2 Cl 4 + 12H 2 0 + CsHsCh => 12C 2 HCl } + 6C0 2 + 15fT + 15CI 
Trichlorobenzene oxidation/ PCE reductive dehalogenation 

-343.92 

-1438 

10.9:1 

0.09:1 

12C 2 HCl 3 + 12 H 2 0 + CsH 3 Cl 3 => 12C 2 H 2 Cl 2 + 6C0 2 + 15fF + 15Ct 
Trichlorobenzene oxidation/ TCE reductive dehalogenation 

-336.96 

-1408 

8.67:1 

0.12:1 

12C 2 H 2 Cl 2 + 12H20+CsH 3 Cl } => 12C 2 H 3 Cl +6C0 2 + 15fF+ 15CI 
Trichlorobenzene oxidation/ cis-DCE reductive dehalogenation 

-279.58 

-1169 

6.40:1 

0.16:1 

12C 2 H 3 Cl +12H 2 0 +C 6 H 3 Cl 3 => 12C 2 H 4 + 6 CO : + 15IT + 15Ct 

Trichlorobenzene oxidation/ Vinyl chloride reductive dehalogenation 

-280.78 

-1174 

4.13:1 

0.24:1 


B3-57 
































Table B. 3 .7 Continued. 


Coupled Tetrachlorobenzene Oxidation Reactions 

AG° f 

(kcal/ mole) 

AG° f 

(kJ/ mole) 

Stoichiometric Mass 
Ratio of Electron 
Acceptor or Metabolic 
Byproduct to Primary 
Substrate 

Mass of Primary 
Substrate Utilized per 
Mass of Electron 
Acceptor Utilized or 
Metabolic Byproduct 
Produced 

5.50: + H 2 0 + C(ti 2 CU => 6C0 2 + 4 ft+ 4Cl 
Tetrachlorobenzene oxidation /aerobic respiration 

-639.10 

-2671 

0.82:1 

1.22:1 

4.4NO}' + 0.4 ft + C 6 H 2 CI 4 => 6 CO 2 + 1.2H 2 0 + 2. 2N 2 . g + 4C1 
Tetrachlorobenzen oxidation / denitrification 

-647.90 

-2708 

1.27: 

0.78:1 

llMnO 2 + 18ft + C 6 H 2 CU => 6 CO 2 + 10H 2 O + llMn 2+ + 4CI 
Tetrachlorobenzenoxidation / manganese reduction 

-639.10 

-2671 

4.47:1 

0.22:1 

22Fe(OHUn + 40ft + C*H,CL => 6 CO , + 54H,0 + 22Fe 2+ + 4CI 

Tetrachlorobenzen oxidation / iron reduction 

-489.50 

-2046 

11.0:1 

0.09:1 

2.75S0 4 2 ' + 1.75ft + H 2 0 + C 6 tf 2 CU => 6C0 2 + 2.75H 2 S° + 4Ct 
Tetrachlorobenzen oxidation / sulfate reduction 

-168.96 

-706.3 

1.23:1 

0.81:1 

6 . 5 H 2 O + C 6 H 2 CI 4 => 3.25C0 2 + 2.75CH 4 + 4ft + 4CI 

Tetrachlorobenzen oxidation / methanogenesis 

-102.52 

-428.5 

0.19:1 

5.19:1 

IIC 2 H 2 CI 4 + 12H : 0 + CtFhCU => 1 lC 2 ff}Cl 3 + 6C0 2 + 15ft + 15CI 
Tetrachlorobenzen oxidation/ PCA reductive dehalogenation 

-287.01 

-1200 

8.53:1 

0.12:1 

IIC 2 H 1 CU + 12H : 0 + => IIC 2 H 4 CI 2 + 6 CO 2 + 15ft + 15Cl 

Tetrachlorobenzen oxidation/ TCA reductive dehalogenation 

-323.64 

-1353 

6.79:1 

0.15:1 

IIC 2 H 4 CI 2 + 1211 2 0 + C(tf 2 Cl 4 => IIC 2 H 5 CI + 6 CO 2 + 15ft + 15Cl 
Tetrachlorobenzen oxidation/DCA reductive dehalogenation 

-297.79 

-1392 

5.04:1 

0.20:1 

IIC 2 CI 4 + 12H 2 0 + C 6 ff 2 Cl 4 => 11C 2 HCl 3 + 6 CO 2 + 15 ft + 15 Cl 
Tetrachlorobenzen oxidation/ PCE reductive dehalogenation 

-313.3 

-1310 

8.43:1 

0.12:1 

IIC 2 HCI 3 + I 2 H 2 O + C 6 H 2 CU => IIC 2 H 2 CI 2 + 6 CO 2 + 15ft + 15Cl 
Tetrachlorobenzen oxidation/ TCE reductive dehalogenation 

-307.03 

-1283 

6.68:1 

0.15:1 

IIC 2 H 2 CI 2 + 12H 2 0 + CtHiCU => UC 2 H 3 CI + 6 CO 2 + 15ft + 15Cl 
Tetrachlorobenzen oxidation/ cis-DCE reductive dehalogenation 

-254.67 

-1065 

4.93:1 

0.20:1 

IIC 2 H 3 CI + 12H 2 0 + C 6 H 2 Cl 4 => IIC 2 H 4 + 6C0 2 + 15ft + 15CI 
Tetrachlorobenzen oxidation/ Vinyl chloride reductive dehalogenation 

-255.77 

-1069 

3.19:1 

0.31:1 


B3-58 


























Anthropogenic Electron 
Acceptors 


PCE Reduction -1500 

I 

TCE Reduction -1465 

I 

cis 1,2-DCE R eductiofil 66 


N atu ral E lectron 
Acceptors AG‘ 


Aerobic Respiration-3202 

\ 

Denitrification -3245 

I 

Manganese (IV) -3202 
R eduction 

\ 

Iron (III) Reduction -2343 




Sulfate Reduction 


\ 

M ethanogenesis 


-514 

-1 36 


* For Benzene Oxidation, kJ/mole 


Figure B.3.4 Expected sequence of microbially-mediated redox reactions and Gibbs free energy of 
reaction. 


B.3.6 ONE-DIMENSIONAL ADVECTION-DISPERSION EQUATION WITH 
RETARDATION AND BIODEGRADATION 

The advection-dispersion equation is obtained by adding a biodegradation term to 
equation B.2.20. In one dimension, this is expressed as: 

dC D x d 2 C v x dC _ 
dt R dx 2 R dx 


eq. B.3.2 


Where: 

v x = average linear ground-water velocity [L/T] 

R = coefficient of retardation 
C = contaminant concentration [M/L 3 ] 

D x = hydrodynamic dispersion [L 2 /T] 
t = time [T] 

jc = distance along flow path [L] 

X = first-order biodegradation decay rate [T 1 ] 

This equation considers advection, hydrodynamic dispersion, sorption (retardation), and biodeg¬ 
radation. First-order rate constants are appropriate for iron (Ill)-reducing, sulfate-reducing, and 
methanogenic conditions. They are not appropriate under aerobic or denitrifying conditions. 


B3-59 







SECTION B-4 

DESTRUCTIVE ATTENUATION MECHANISMS - ABIOTIC 

Chlorinated solvents dissolved in ground water may also be degraded by abiotic mechanisms, 
although the reactions are typically not complete and often result in the formation of an intermediate 
that may be at least as toxic as the original contaminant. The most common reactions affecting 
chlorinated compounds are hydrolysis (a substitution reaction) and dehydrohalogenation (an elimina¬ 
tion reaction). Other possible reactions include oxidation and reduction reactions. Butler and Barker 
(1996) note that no abiotic oxidation reactions involving typical halogenated solvents have been 
reported in the literature. They also note that reduction reactions (which include hydrogenolysis and 
dihaloelimination) are commonly microbially mediated, although some abiotic reduction reactions 
have been observed. 

As Butler and Barker (1996) note, attributing changes in either the presence or absence of 
halogenated solvents or the concentrations of halogenated solvents to abiotic processes is usually 
difficult. For example, microbial activity is generally required to produce reducing conditions that 
favor reductive dehalogenation. If such activity is taking place, chlorinated solvents may be under¬ 
going both biotic and abiotic degradation, and discerning the relative contribution of each mecha¬ 
nism on the field scale, if possible, would be very difficult. As another example, Butler and Barker 
(1996) note that to substantiate that hydrolysis is occurring, the presence of non-halogenated break¬ 
down products such as acids and alcohols should be established. In general, these products are more 
easily biodegraded than their parent compounds and can be difficult to detect. Field evidence of this 
nature has yet to be collected to demonstrate hydrolysis of halogenated solvents (Butler and Barker, 
1996). 

Given the difficulties of demonstrating abiotic degradation on the field scale, it may not be 
practical to demonstrate that such processes are occurring and to quantitatively evaluate the contribu¬ 
tions of those reactions (i.e., separately from biotic processes). If biodegradation is occurring at a 
site, the loss of contaminant mass due to that process may dwarf the mass lost to abiotic reactions, 
ruling out a cost-effective evaluation of abiotic degradation. However, while the rates of abiotic 
degradation may be slow relative to biotic mechanisms, the contribution of these mechanisms may 
still play a significant role in natural attenuation, depending on site conditions (e.g., a site with a 
slow solute transport velocity or a long distance to the nearest receptor). Vogel (1994) describes data 
patterns that may result from varying combinations of biotic and abiotic degradation of chlorinated 
solvents. Moreover, because some of the by-products of these reactions are chlorinated compounds 
that may be more easily or less easily degraded than the parent, the contributions of abiotic mecha¬ 
nisms may need to be considered when evaluating analytical data from a site. 

B.4.1 HYDROLYSIS AND DEHYDROHALOGENATION 

As discussed by Butler and Barker (1996), hydrolysis and dehydrohalogenation reactions are the 
most thoroughly studied abiotic attenuation mechanisms. In general, the rates of these reactions are 
often quite slow within the range of normal ground-water temperatures, with half-lives of days to 
centuries (Vogel et al., 1987; Vogel, 1994). Therefore, most information about the rates of these 
reactions is extrapolated from experiments run at higher temperatures so that the experiments could 
be performed within a practical time frame. 

B.4.1.1 Hydrolysis 

Hydrolysis is a substitution reaction in which an organic molecule reacts with water or a com¬ 
ponent ion of water, and a halogen substituent is replaced with a hydroxyl (OH ) group. The hy¬ 
droxyl substitution typically occurs at the halogenated carbon. This leads initially to the production 
of alcohols. If the alcohols are halogenated, additional hydrolysis to acids or diols may occur. Also, 


B4-60 


the addition of a hydroxyl group to a parent molecule may make the daughter product more suscep¬ 
tible to biodegradation, as well as more soluble (Neely, 1985). Non-alcohol products have also been 
reported by Vogel et al. (1987) and Jeffers et al. (1989), but they are apparently products of compet¬ 
ing dehydrohalogenation reactions. 

The likelihood that a halogenated solvent will undergo hydrolysis depends in part on the num¬ 
ber of halogen substituents. More halogen substituents on a compound will decrease the chance for 
hydrolysis reactions to occur (Vogel et al ., 1987), and will therefore decrease the rate of the reaction. 
In addition, bromine substituents are more susceptible to hydrolysis than chlorine substituents (Vogel 
et al., 1987). 1,2-Dibromoethane is one compound that is subject to significant hydrolysis reactions 
under natural conditions. Locations of the halogen substituent on the carbon chain may also have 
some effect on the rate of reaction. The rate also may increase with increasing pH; however, a rate 
dependence upon pH is typically not observed below a pH of 11 (Mabey and Mill, 1978; Vogel and 
Reinhard, 1986). Rates of hydrolysis may also be increased by the presence of clays, which can act 
as catalysts (Vogel et al., 1987). Hydrolysis rates can generally be described using first-order kinet¬ 
ics, particularly in solutions in which water is the dominant nucleophile (Vogel et al., 1987). How¬ 
ever, this oversimplifies what is typically a much more complicated relationship (Neely, 1985). As 
noted in the introduction to this Appendix, reported rates of environmentally significant hydrolysis 
reactions involving chlorinated solvents are typically the result of extrapolation from experiments 
performed at higher temperatures (Mabey and Mill, 1978; Vogel, 1994). 

Hydrolysis of chlorinated methanes and ethanes has been well-demonstrated in the literature. 
Vogel (1994) reports that monohalogenated alkanes have half-lives on the order of days to months, 
while polychlorinated methanes and ethanes have half-lives that may range up to thousands of years 
for carbon tetrachloride. As the number of chlorine atoms increases, dehydrohalogenation may 
become more important (Jeffers et al., 1989). Butler and Barker (1996) note that chlorinated ethenes 
do not undergo significant hydrolysis reactions (i.e., the rates are slow). Butler and Barker also 
reported that they were unable to find any studies on hydrolysis of vinyl chloride. A listing of half- 
lives for abiotic hydrolysis and dehydrohalogenation of some chlorinated solvents is presented on 
Table B.4.1. Note that no distinctions are made in the table as to which mechanism is operating; this 
is consistent with the references from which the table has been derived (Vogel et al., 1987; Butler 
and Barker, 1996). 

One common chlorinated solvent for which abiotic transformations have been well-studied is 
1,1,1 -TCA. 1,1,1 -TCA may be abiotically transformed to acetic acid through a series of substitution 
reactions, including hydrolysis. In addition, 1,1,1-TCA may be reductively dehalogenated to form 
1,1- DC A) and then chloroethane (CA), which is then hydrolyzed to ethanol (Vogel and McCarty, 
1987) or dehydrohalogenated to vinyl chloride (Jeffers et al., 1989). Rates of these reactions have 
been studied by several parties, and these rates are summarized in Table B.4.1. 

B.4.1.2 Dehydrohalogenation 

Dehydrohalogenation is an elimination reaction involving halogenated alkanes in which a 
halogen is removed from one carbon atom, followed by the subsequent removal of a hydrogen atom 
from an adjacent carbon atom. In this two-step reaction, an alkene is produced. Although the 
oxidation state of the compound decreases due to the removal of a halogen, the loss of a hydrogen 
atom increases it. This results in no external electron transfer, and there is no net change in the 
oxidation state of the reacting molecule (Vogel et al., 1987). Contrary to the patterns observed for 
hydrolysis, the likelihood that dehydrohalogenation will occur increases with the number of halogen 
substituents. It has been suggested that under normal environmental conditions, monohalogenated 
aliphatics apparently do not undergo dehydrohalogenation, and these reactions are apparently not 
likely to occur (March, 1985; Vogel et al., 1987). However, Jeffers et al. (1989) report on the 


B4-61 


dehydrohalogenation of CA to VC. Polychlorinated alkanes have been observed to undergo 
dehydrohalogenation under normal conditions and extremely basic conditions (Vogel et al., 1987). 
As with hydrolysis, bromine substituents are more reactive with respect to dehydrohalogenation. 

Table B.4.1 Approximate Half-Lives of Abiotic Hydrolysis andDehydrohalogenation Reactions Involving 
Chlorinated Solvents 


Compound 

Half-Life (years) 

Products 

Chloromethane 

no data 


Methylene Chloride 
(Dichloromethane) 

704^ 


Trichloromethane 

3500 a , 1800 b 


(Chloroform) 



Carbon Tetrachloride 

41 b 


Chloroethane 

0.12 c 

ethanol 

1,1 -Dichloroethane 

61 b 


1,2-Dichloroethane 

72 b 


1,1,1 -Trichloroethane 

1.7 a , l.l b 

acetic acid 


2.5 d 

1,1-DCE 

1,1,2-Trichloroethane 

140\ 170 a 

1,1 -DCE 

1,1,1,2-Tetrachloroethane 

47 b , 380 a 

TCE 

1,1,2,2-Tetrachloroethane 

0.3 e 

1,1,2-TCA 


0.4 b , 0.8 a 

TCE 

Tetrachloroethene 

0.7 f *, 1.3 x 10 6b 


Trichloroethene 

0.7 f *, 1.3 x 10 6b 


1,1 -Dichloroethene 

1.2 x 10 8b 


1,2-Dichloroethene 

2.1 xl0 10b 



s From Mabey and Mill, 1978 
b From Jeffers et al., 1989 
c From Vogel et al., 1987 
d From Vogel and McCarty, 1987 

* From Cooper et al., 1987 
f From Dilling et al., 1975 

* Butler and Barker (1996) indicate that these values may reflect experimental difficulties 
and that the longer half-life [as calculated by Jeffers et al. (1989)] should be used. 


Dehydrohalogenation rates may also be approximated using pseudo-first-order kinetics. Once 
again, this is not truly a first-order reaction, but such approximations have been used in the literature 
to quantify the reaction rates. The rates will not only depend upon the number and types of halogen 
substituent, but also on the hydroxide ion concentration. Under normal pH conditions (i.e., near a 


B4-62 




















pH of 7), interaction with water (acting as a weak base) may become more important (Vogel et al., 
1987). Transformation rates for dehydrohalogenation reactions is presented in Table B.4.1. 1,1,1- 
TCA is also known to undergo dehydrohalogenation (Vogel and McCarty, 1987). In this case, TCA 
is transformed to 1,1-DCE, which is then reductively dehalogenated to VC. The VC is then either 
reductively dehalogenated to ethene or consumed as a substrate in an aerobic reaction and converted 
to C0 2 . In a laboratory study, Vogel and McCarty (1987) reported that the abiotic conversion of 
1,1,1 -TCA to 1,1 -DCE has a rate constant of about 0.04 year 1 . It was noted that this result was 
longer than indicated in previous studies, but that experimental methods differed. Jeffers et al. 
(1989) reported on several other dehydrohalogenation reactions; in addition to 1,1,1-TCA and 1,1,2- 
TCA both degrading to 1,1-DCE, the tetrachloroethanes and pentachloroethanes degrade to TCE and 
PCE, respectively. Rates of these reactions are included in Table B.4.1. As noted previously, Jeffers 
et al. (1989) also report that CA may degrade to VC, but no information on rates was encountered 
during the literature search for this Appendix. 

B.4.2 REDUCTION REACTIONS 

Two abiotic reductive dechlorination reactions that may operate in the subsurface are 
hydrogenolysis and dihaloelimination. Hydrogenolysis is the simple replacement of a chlorine (or 
another halogen) by a hydrogen, while dihaloelimination is the removal of two chlorines (or other 
halogens) accompanied by the formation of a double carbon-carbon bond. Butler and Barker (1996) 
review work by Criddle et al. (1986), Jafvert and Wolfe (1987), Reinhard et al. (1990), and Acton 
(1990) and this review suggests that while these reactions are thermodynamically possible under 
reducing conditions, they often do not take place in the absence of biological activity, even if such 
activity is only indirectly responsible for the reaction. While not involved in a manner similar to that 
for cometabolism, microbes may produce reductants that facilitate such reactions in conjunction with 
minerals in the aquifer matrix, as has been suggested by work utilizing aquifer material from the 
Borden test site (Reinhard et al ., 1990). Moreover, the reducing conditions necessary to produce 
such reactions are most often created as a result of microbial activity. It is therefore not clear if some 
of these reactions are truly abiotic, or if because of their reliance on microbial activity to produce 
reducing conditions or reactants, they should be considered to be a form of cometabolism. 

In some cases, truly abiotic reductive dechlorination has been observed; however, the conditions 
that favor such reactions may not occur naturally. For example, Gillham and O’Hannesin (1994) 
describe reductive dehalogenation of chlorinated aliphatics using zero-valent iron, in which the iron 
serves as an electron donor in an electrochemical reaction. However, this is not a natural process. 
Wang and Tan (1990) reported reduction of TCE to ethene and carbon tetrachloride to methane 
during a platinum-catalyzed reaction between elemental magnesium and water. Given that the 
metals involved in these reactions are unlikely to occur naturally in the reduced forms used in the 
aforementioned work, such processes are not likely to contribute to natural attenuation of chlorinated 
solvents. 


B4-63 





■ ' 
































































APPENDIX C 


DATA INTERPRETATION and CALCULATIONS 


TABLE OF CONTENTS - APPENDIX C 


C-1 INTRODUCTION.C1 - 5 

C-2 PREPARATION OF GEOLOGIC BORING LOGS, 

HYDROGEOLOGIC SECTIONS, AND MAPS.C2-6 

C.2.1 PREPARATION OF LITHOLOGIC LOGS.C2-6 

C.2.2 PREPARATION OF HYDROGEOLOGIC SECTIONS.C2-7 

C.2.3 REVIEW OF TOPOGRAPHIC MAPS AND PREPARATION OF 

POTENTIOMETRIC SURFACE MAPS AND FLOW NETS.C2-7 

C.2.3.1 Review of Topographic Maps.C2-7 

C.2.3.2 Preparation of Potentiometric Surface Maps.C2-7 

C.2.3.3 Preparation of Flow Nets.C2-9 

C.2.3.4 Preparation of Contaminant Isopach Maps.C2-9 

C.2.3.5 Preparation of Contaminant and Daughter Product Isopleth Maps.C2-14 

C.2.3.6 Preparation of Electron Donor, Inorganic Electron Acceptor, and 

Metabolic By-product Contour (Isopleth) Maps.C2-15 

C-3 NATURAL ATTENUATION CALCULATIONS.C3-18 

C.3.1 CALCULATING HYDRAULIC PARAMETERS.C3-18 

C.3.1.1 Hydraulic Conductivity.C3 -18 

C.3.1.2 Transmissivity.C3-20 

C.3.1.3 Hydraulic Head and Gradient.C3-20 

C.3.1.4 Total Porosity (n) and Effective Porosity (n e ).C3-23 

C.3.1.5 Linear Ground-water Flow Velocity (Seepage or Advective Velocity).C3-24 

C.3.1.6 Coefficient of Retardation and Retarded Contaminant Transport Velocity.C3-25 

C.3.2 CONTAMINANT SOURCE TERM CALCULATIONS.C3-28 

C.3.2.1 Direct Measurement of Dissolved Contaminant Concentrations in 

Ground Water in Contact with NAPL.C3-31 

C.3.2.2 Equilibrium Partitioning Calculations.C3-32 

C.3.2.3 Mass Flux Calculations.•..C3-33 

C.3.3 CONFIRMING AND QUANTIFYING BIODEGRADATION.C3-37 

C.3.3.1 IsoplethMaps.C3-37 

C.3.3.2 Data Set Normalization.C3-37 

C.3.3.3 Calculating Biodegradation Rates.C3-41 

C.3.4 DESIGN, IMPLEMENTATION, AND INTERPRETATION OF 

MICROCOSM STUDIES.C3-49 

C.3.4.1 Overview.C3-49 

C.3.4.2 When to Use Microcosms.C3-50 

C.3.4.3 Application of Microcosms.C3-50 

C.3.4.4 Selecting Material for Study.C3-50 

C.3.4.5 Geochemical Characterization of the Site.C3-51 

C.3.4.6 Microcosm Construction.C3-54 

C.3.4.7 Microcosm Interpretation.C3-54 

C.3.4.8 The Tibbetts Road Case Study.C3-55 

C.3.4.9 Summary.C3-58 


Cl-2 







































FIGURES 


No. Title Page 

C.2.1 Example hydrogeologic section.C2-6 

C.2.2 Example ground-water elevation map.C2-8 

C.2.3 Example flow net.C2-9 

C.2.4 Example mobile LNAPL isopach (A) and contaminant isopleth (B) maps.C2-10 

* C.2.5 Measured (apparent) versus actual LNAPL thickness.C2-11 

C.2.6 Type curve for LNAPL baildown test.C2-14 

C.2.7 Example isopleth maps of contaminants and soluble electron acceptors.C2-16 

C.2.8 Example isopleth maps of contaminants and metabolic by-products.C2-17 

C.3.1 Range of hydraulic conductivity values.C3-18 

C.3.2 Hydraulic head.C3-21 

C.3.3 Ground-water Elevation Map.C3-23 

C.3.4 Location of sampling points at the St. Joseph, Michigan NPL site.C3-25 

C.3.5 Field rate constants for TCE as reported in literature.C3-43 

C.3.6 Field rate constants for PCE as reported in literature.C3-43 

C.3.7 Field rate constants for Vinyl Chloride as reported in literature.C3-44 

C.3.8 Exponential regression of TCE concentration on time of 

travel along flow path.C3-46 

C.3.9 Regression of the TCE concentration on distance along flow path.C3-48 

C.3.10 Tibbetts Road study site.C3-49 

C.3.11 TCE microcosm results.C3-56 

C.3.12 Benzene microcosm results.C3-56 

C.3.13 Toluene microcosm results.C3-57 


Cl-3 























TABLES 


No. 


Title 


Page 


C.2.1 Typical Values for dr .C2-12 

C.2.2 Surface Tensions for Various Compounds.C2-13 

C.2.3 Results of Example Baildown Test.C2-15 

C.3.1 Representative Values of Hydraulic Conductivity for Various 

Sediments and Rocks.C3-19 

C.3.2 Representative Values of Dry Bulk Density, Total Porosity, and 

Effective Porosity for Common Aquifer Matrix Materials...C3-24 

C.3.3 Representative Values of Total Organic Carbon 

for Common Sediments.C3-27 

C.3.4 Example Retardation Calculations for Select Compounds.C3-28 

C.3.5 Attenuation of Chlorinated Ethenes and Chloride Downgradient 

of the Source of TCE in the West Plume at the St. Joseph, Michigan, NPL Site.C3-40 

C.3.6 Use of the Attenuation of a Tracer to Correct the Concentration of TCE 
Downgradient of the Source of TCE in the West Plume at the 

St. Joseph, Michigan NPL Site.C3-41 

C.3.7 Geochemical Parameters Important to Microcosm Studies.C3-53 

C.3.8 Contaminants and Daughter Products.C3-53 

C.3.9 Concentrations of TCE, Benzene, and Toluene in the 

Tibbetts Road Microcosms.C3-58 

C.3.10 First-order Rate Constants for Removal of TCE, Benzene, and Toluene 

in the Tibbetts Road Microcosms.C3-59 

C.3.11 Concentrations of Contaminants and Metabolic By-products in Monitoring 

Wells along Segments in the Plume used to Estimate Field-scale Rate Constants.C3-59 

C.3.12 Comparison of First-order Rate for Contaminant Attenuation in Segments 

of the Tibbetts Road Plume.. C3-60 

C.3.13 Comparison of First-order Rate Constants in a Microcosm Study, 

and in the Field at the Tibbetts Road NPL Site.C3-60 


Cl-4 


















SECTION C-l 
INTRODUCTION 

Successful documentation of natural attenuation requires interpretation of site-specific data to 
define the ground-water flow system, refine the conceptual model, quantify rates of contaminant 
attenuation, and model the fate and transport of dissolved contaminants. Tasks to be completed 
include preparation of lithologic logs, hydrogeologic sections, potentiometric surface maps and flow 
nets, contaminant isopach and isopleth maps, electron acceptor and metabolic byproduct isopleth 
maps, and calculation of hydraulic parameters, retardation coefficients, and biodegradation rate 
constants. The rate and amount of partitioning of organic compounds from mobile and residual 
nonaqueous-phase liquid (NAPL) into ground water should also be determined to allow estimation 
of a source term. Completion of these tasks permits refinement of the conceptual model and is 
necessary to successfully support remediation by natural attenuation. 

This appendix consists of three sections, including this introduction. Section C-2 discusses 
preparation of geologic boring logs, hydrogeologic sections, and maps. Section C-3 covers natural 
attenuation calculations, including hydraulic parameter calculations, contaminant source term 
calculations, confirming and quantifying biodegradation, and designing, implementing, and 
interpreting microcosm studies. 


Cl-5 


SECTION C-2 

PREPARATION OF GEOLOGIC BORING LOGS, HYDROGEOLOGIC 

SECTIONS, AND MAPS 

The first step after completion of site characterization field activities is to prepare geologic 
boring logs, hydrogeologic sections, water table elevation (or potentiometric surface) maps, flow 
nets, and maps depicting contaminant concentrations, electron acceptor and metabolic byproduct 
concentrations, and mobile NAPL thickness. The construction of these items is discussed in the 
following sections. 

C.2.1 PREPARATION OF LITHOLOGIC LOGS 

Lithologic logs should be prepared using field data. Whenever possible, these logs should 
contain descriptions of the aquifer matrix, including relative density, color, major textural constitu¬ 
ents, minor constituents, porosity, relative moisture content, plasticity of fines, cohesiveness, grain 
size, structure or stratification, relative permeability, and any significant observations such as visible 
fuel or fuel odor. It is also important to correlate the results of volatile organic compound (VOC) 
screening using headspace vapor analysis with depth intervals of geologic materials. The depth of 
lithologic contacts and/or significant textural changes should be recorded to the nearest 0.1 foot. 

This resolution is necessary because preferential flow and contaminant transport pathways may be 
limited to stratigraphic units less than 6 inches thick. 

C.2.2 PREPARATION OF HYDROGEOLOGIC SECTIONS 

Lithologic logs should be used in conjunction with water level data to prepare a minimum of 
two hydrogeologic sections for the site. One section should be oriented parallel to the direction of 
ground-water flow, and one section should be oriented perpendicular to the direction of ground-water 
flow. Both sections should be drawn to scale. Hydrogeologic sections are an integral part of the 
conceptual model and are useful in identifying preferential contaminant migration pathways and in 
modeling the site. 

At a minimum, hydrogeologic sections should contain information on the relationships between 
hydrostratigraphic units at the site, including the location and distribution of transmissive vs. non¬ 
transmissive units, the location of the water table relative to these units, and the location(s) of the 
contaminant source(s). Figure C.2.1 is an example of a completed hydrogeologic section. 



Figure C.2.1 Example hydrogeologic section. 


C2-6 

























C.2.3 REVIEW OF TOPOGRAPHIC MAPS AND PREPARATION OF POTENTIOMETRIC 
SURFACE MAPS AND FLOW NETS 

Determining the direction of ground-water flow and the magnitude of hydraulic gradients is 
important because these parameters influence the direction and rate of contaminant migration. 
Ground-water flow directions are represented by a three-dimensional set of equipotential lines and 
orthogonal flow lines. If a plan view (potentiometric surface, or water table elevation, map) or a 
two-dimensional cross-section is drawn to represent a flow system, the resultant equipotential lines 
and flow lines constitute a flow net. A flow net can be used to determine the distribution of hydrau¬ 
lic head, the ground-water velocity distribution, ground-water and solute flow paths and flow rates, 
and the general flow pattern in a ground-water system. 

C.2.3.1 Review of Topographic Maps 

Ground-water flow is strongly influenced by the locations of ground-water divides and by 
recharge from and discharge to surface waterbodies such as rivers, streams, lakes, and wetlands. 
Topographic highs generally represent divergent flow boundaries (divergent ground-water divide), 
and topographic lows such as valleys or drainage basins typically represent convergent flow bound¬ 
aries (convergent ground-water divide). In addition, the configuration of the water table is typically a 
subtle reflection of the surface topography in the area. However, topography is not always indicative 
of subsurface flow patterns and should not be depended upon unless confirmed by head data. In 
order to place the local hydrogeologic flow system within the context of the regional hydrogeologic 
flow system, it is important to have an understanding of the local and regional topography. Included 
in this must be knowledge of the locations of natural and manmade surface water bodies. This 
information can generally be gained from topographic maps published by the United States Geologi¬ 
cal Survey. 

C.2.3.2 Preparation of Potentiometric Surface Maps 

A potentiometric surface map is a two-dimensional graphical representation of equipotential 
lines shown in plan view. Water table elevation maps are potentiometric surface maps drawn for 
water table (unconfined) aquifers. Potentiometric surface maps for water table aquifers show where 
planes of equal potential intersect the water table. A potentiometric surface map should be prepared 
from water level measurements and surveyor’s data. These maps are used to estimate the direction 
of plume migration and to calculate hydraulic gradients. To document seasonal variations in ground- 
water flow, separate potentiometric surface maps should be prepared using quarterly water level 
measurements taken over a period of at least 1 year. 

The data used to develop the potentiometric surface map should be water level elevation data 
(elevation relative to mean sea level) from piezometers/wells screened in the same relative position 
within the same hydrogeologic unit. For example, wells that are screened at the water table can be 
used for the same potentiometric surface map. Wells screened in different hydrogeologic units or at 
different relative positions within the same water table aquifer cannot be used to prepare a potentio¬ 
metric surface map. Where possible, a potentiometric surface map should be prepared for each 
hydrogeologic unit at the site. In recharge areas, wells screened at various elevations cannot all be 
used to prepare the same potentiometric surface map because of strong downward vertical gradients. 
Likewise, wells screened at various elevations in discharge areas such as near streams, lakes, or 
springs, should not all be used because of the strong upward vertical gradients. 

When preparing a potentiometric surface map, the locations of system boundaries should be 
kept in mind; particularly the site features that tend to offset the shape of the contours on the map. 
Such features include topographic divides, surface water bodies, and pumping wells. 

In addition to, and separately from, preparation of a potentiometric surface map, water level 
measurements from wells screened at different depths can be used to determine any vertical hydrau- 


C2-7 


lie gradients. It is important to have a good understanding of vertical hydraulic gradients because 
they may have a profound influence on contaminant migration. 

In areas with measurable mobile LNAPL, a correction must be made for the water table deflec¬ 
tion caused by the LNAPL. The following relationship, based on Archimedes’ Principle, provides a 
correction factor that allows the water table elevation to be adjusted for the effect of floating 
LNAPL. 


CDTW=MDIW(PT) ea C 2 1 

A, 

Where: 

CDTW = corrected depth to water [L] 

MDTW - measured depth to water [L] 

Ptnapi ~ density of the LNAPL [M/L 3 ] 

p H = density of the water, generally 1.0 [M/L 3 ] 

PT = measured LNAPL thickness [L] 

Using the corrected depth to water, the corrected ground-water elevation, CGWE, is given by: 

CGWE = Datum Elevation - CDTW eq. C.2.2 

Corrected ground-water elevations should be used for potentiometric surface map preparation. 

Figure C.2.2 is an example of a ground-water elevation map for an unconfined aquifer. Water table 
elevation data used to prepare this map were taken from wells screened across the water table. 



Figure C.2.2 Example ground-water elevation map. 


C2-8 















































C.2.3.3 Preparation of Flow Nets 

Where an adequate three-dimensional database is available, flow nets can be constructed to 
facilitate the interpretation of the total hydraulic head distribution in the aquifer. This will help 
determine potential solute migration pathways. The simplest ground-water flow system is one that is 
homogeneous and isotropic. This type of hydrogeologic setting serves as a simple basis for describ¬ 
ing the basic rules of flow net construction, despite the fact that homogeneous, isotropic media rarely 
occur in nature. Regardless of the type of geologic media, the basic rules of flow net construction 
must be applied, and necessary modifications must be made throughout the procedure to account for 
aquifer heterogeneity or anisotropic conditions. Water level data for flow net construction should 
come from multiple sets of nested wells (two or more wells at the same location) at various depths in 
the aquifer. The fundamental rules of flow net construction and the important properties of flow nets 
are summarized as follows: 

• Flow lines and equipotential lines intersect at 90-degree angles if the permeability is 
isotropic; 

• The geometric figures formed by the intersection of flow lines and equipotential lines must 
approximate squares or rectangles; 

» Equipotential lines must meet impermeable boundaries at right angles (impermeable 
boundaries are flow lines); and 

• Equipotential lines must be parallel to constant-head boundaries (constant-head bound¬ 
aries are equipotential lines). 

Trial-and-error sketching is generally used to construct a flow net. Flow net sketching can be 
sufficiently accurate if constructed according to the basic rules outlined above. A relatively small 
number of flow lines (three to five) generally are sufficient to adequately characterize flow condi¬ 
tions. Flow nets should be superimposed on the hydrogeologic sections. Figure C.2.3 is an example 
of a completed flow net. This figure shows ground-water flow patterns in both recharge and dis¬ 
charge areas. 

C.2.3.4 Preparation of Contaminant Isopach Maps 

If NAPL is present at the site, isopach maps showing the thickness and distribution of NAPL 
should be prepared. Two maps should be prepared: one for mobile NAPL, and one for residual 
NAPL. Such isopach maps allow estimation of the distribution of NAPL in the subsurface and aid in 



—► Flow Line 

Equipotential Line 
10 Total Head (meters) 


Figure C.2.3 Example flow net. 


C2-9 
























fate and transport model development by identifying the boundary of the NAPL. Because of the 
differences between the magnitude of capillary suction in the aquifer matrix and the different surface 
tension properties of fuel and water, LNAPL thickness observations made in monitoring points are 
only an estimate of the actual volume of mobile LNAPL in the aquifer. To determine the actual 
NAPL thickness it is necessary to collect and visually analyze soil samples. LNAPL thickness data 
also should be used to correct for water table deflections caused by the mobile LNAPL. This process 
is described in Section C.2.2.3.2. 

Isopach maps are prepared by first plotting the measured NAPL thickness on a base map pre¬ 
pared using surveyor’s data. Lines of equal NAPL thickness (isopachs) are then drawn and labeled. 
Each data point must be honored during contouring. Figure C.2.4 is an example of a completed 
isopach map. This figure also contains an example of an isopleth map. 

C.2.3.4.1 Relationship Between Apparent and Actual LNAPL Thickness 

It is well documented that LNAPL thickness measurements taken in ground-water monitoring 
wells are not indicative of actual LNAPL thicknesses in the formation (de Pastrovich et al., 1979; 
Blake and Hall, 1984; Hall et al. , 1984; Hughes et al , 1988; Abdul et al., 1989; Testa and 
Paczkowski, 1989; Farr et al., 1990; Kemblowski and Chiang, 1990; Lenhard and Parker, 1990; 
Mercer and Cohen, 1990; Ballestero et al., 1994; Huntley et al. , 1994a). These authors note than the 
measured thickness of LNAPL in a monitoring well is greater than the true LNAPL thickness in the 
aquifer and, according Mercer and Cohen (1990), measured LNAPL thickness in wells is typically 2 
to 10 times greater than the actual LNAPL thickness in the formation. The difference between actual 
and measured LNAPL thickness occurs because mobile LNAPL floating on the water table flows 
into the well (if the top of well screen is above the base of the LNAPL) and depresses the water 
table. Figure C.2.5 is a schematic that illustrates this relationship. The equation for correcting depth 


A) MOBILE LNAPL ISOPACH MAP 


B) CONTAMINANT ISOPLETH MAP 


SOURCE AREA 



Q36 

1.00 


UVft lHOtfvE$(FB) 

UINECTEQJOLUSffl 

1HOOSBB(FE) 

COJCLRlMER/l = 1FCO 


127 TaA.BIEXa>CBMRHlCN{ gij (i 

-1000 - 

CDsCBWiojf gy n 
(CX&flJWSENBOT) 
aMOKNU»A=lXn} 0L n 


O 150 |00 


;! 

T 


Figure C.2.4 Example mobile LNAPL isopach (A) and contaminant isopleth (B) maps. 


C2-10 













































































to ground water caused by LNAPL in the well is given in Section C.2.3.2. Empirical relationships 
relating measured LNAPL thickness to actual LNAPL thickness are presented below. Also presented 
below are test methods that can be used to determine actual LNAPL thickness. There are no estab¬ 
lished methods for determining actual DNAPL volume based on measurements taken in monitoring 
wells. 


LNAPLFraction at or Below 
Residual Saturation 



Zone of LNAPL 
Capillary Rise 


Actual 

LNAPL Thickness 


Zone of Water 
Capillary Rise 


LNAPL 

Fraction Greater Than 
Residual Saturation 


Top of Product 7 ^ 


Apparent 
LNAPL Thickness 


LEGEND Measured Water Table * 

Residual Hydrocarbons 
Free Liquid Hydrocarbons 


Source: Modified from de Pastrovich and others, 1972. 


Figure C.2.5 Measured (apparent) versus actual LNAPL thickness. 


C.2.3.4.2. Empirical Relationships 

There are several empirical methods available to estimate the actual thickness of mobile 
LNAPL in the subsurface based on LNAPL thicknesses measured in a ground-water monitoring 
well. Such empirical relationships are, at best, approximations because many factors influence the 
relationship between measured and apparent LNAPL thickness (Mercer and Cohen, 1990): 

• Capillary fringe height depends on grain size and is hysteretic with fluid level fluctuations. 

• LNAPL can become trapped below the water table as the water table rises and falls. 

• The thickness of LNAPL is ambiguous because the interval of soil containing mobile 
LNAPL is not 100-percent saturated with LNAPL. 

Some empirical methods for determining actual LNAPL thickness are described below. 

Method of de Pastrovich et al. (1979) 

Hampton and Miller (1988) conducted laboratory experiments to examine the relationship 
between the actual thickness of LNAPL in a formation, h^, and that measured in a monitoring well, 
h . Based on their research, Hampton and Miller (1988) suggest using the following relationship 
(developed by de Pastrovich et al., 1979) to estimate LNAPL thickness: 


C2-11 
































eq. C.2.3 


h y ~ 


h m {Pw Plnapl ) 


Pnapl 


Where: 


h f = actual thickness of LNAPL in formation 
/* = measured LNAPL thickness in well 

m 

p v = density of water (1.0 gm/cm 3 for pure water) 
p lnapl - density of LNAPL (See Table C.3.9) 

Method of Kemblowski and Chiang 09901 

Another empirical relationship was proposed by Kemblowski and Chiang (1990) to estimate 
actual LNAPL thickness based on measured LNAPL thickness. This relationship is given by: 


h=H-2.2h 


c 

aw 


dr 


eq. C.2.4 


Where: 

h g = equivalent thickness of LNAPL in the formation (volume of oil per unit area of aquifer, 
divided by porosity) 

H = measured LNAPL thickness in well 

o 

h c aw dr = capillary height of air-water interface assuming water is being displaced by oil 
(typical values are given in Table C.2.1) 

This method assumes equilibrium conditions, water drainage, and oil imbibition. 


Table C.2.1 Typical Values for h c aw dr (Bear, 1972) 


Aquifer Matrix 

K. Jr (cm) 

* ( ft ) 

Coarse Sand 

2-5 

0.066-0.16 

Sand 

12-35 

0.39-1.15 

Fine Sand 

35-70 

1.14-2.30 

Silt 

70-150 

2.30-4.92 

Clay 

>200-400 

>6.56-13.12 


Method of Lenhard and Parker (1990) 

Another empirical relationship was proposed by Lenhard and Parker (1990) to estimate actual 
LNAPL thickness based on measured LNAPL thickness. This relationship is given by: 

n __ Pro fiao _ 

’"Aa-/L(i-a) eqC: 

Where: 

D = actual thickness of LNAPL in formation 

o 

H - measured LNAPL thickness in well 

o 

p w - specific gravity of LNAPL (density of oil/density of water) 

P ao = "^“Air-oil scaling factor 


C2-12 





















aw 

p ow = “““Oil-water scaling factor 

U 0W' 

= surface tension of uncontaminated water (72.75 dynes/cm @ 20°C) 

G ao = surface tension of LNAPL [25 dynes/cm @ 20°C for JP-4, Table C.2.2] 
a w , = - o ao = interfacial tension between water and LNAPL (47.75dynes/cm @ 20°C) 

It is important to note that this method includes the capillary thickness of the hydrocarbon, and is, 
therefore, likely to be an overestimate. 

Table C.2.2 Surface Tensions for Various Compounds 


Compound 

Surface Tension (a), 20°C (dyne/cm) 

JP-4 

25 a/ 

Gasoline 

19-23^ 

Pure Water 

72.75 w 


a/ Martel (1987). 
b/CRC Handbook (1956). 

C.2.3.4.3. LNAPL Baildown Test 

The LNAPL baildown test is applicable in areas where the hydrocarbon/water interface is below 
the potentiometric surface, and the recharge rate of hydrocarbon into the well is slow (Hughes et al., 
1988). 

Baildown Test Procedure (from Hughes et al ., 1988): 

1) Gauge the well and calculate the corrected potentiometric surface elevation using equations 
C.2.1 and C.2.2. 

2) Rapidly bail the hydrocarbon from the well. 

3) Gauge the well again, and if the thickness of the hydrocarbon is acceptable (0.1 to 1 foot), 
calculate the potentiometric surface elevation. The potentiometric surface elevation thus 
calculated should be within 0.005 foot of the value calculated in step 1. If it is, then continue 
to step 4; if it is not, repeat steps 2 and 3. 

4) Record the top of the LNAPL surface in the well as it recharges until the well is fully re¬ 
charged. 

5) Plot the elevation of the top of LNAPL in the well vs. time since bailing ceased. 

6) The true thickness of the mobile LNAPL layer (T f ) is the distance from the inflection point to 
the top of the hydrocarbon under static conditions (Figure C.2.6). Thus, T f is picked directly 
off the plot. Table C.2.3 is an example of the results of this procedure. 


C2-13 










Figure C.2.6 Type curve for LNAPL baildown test. 


Table C.2.3 Results of Example Baildown Test (Modified from Hughes et al., 1988) 


Well 

T w 

T f 

Exaggeration (T w /Tf) 


(ft) 1 ' 

(ft) 


ROW-143 

4.97 

0.61 

8.1:1 

ROW-189 

12.5 

0.29 

43.0:1 

ROW-129 

0.94 

o.o b ' 

N/A 


a/ T w = LNAPL thickness initially measured in the well, if LNAPL thickness that is actually 
mobile 

b/ Capillary oil only 

Hughes et al. (1988) also present a recharge method that involves pumping the mobile LNAPL until 
steady-state conditions are achieved, and then letting the well fully recharge. 

C.2.3.5 Preparation of Contaminant and Daughter Product Isopieth Maps 

Isopleth maps should be prepared for all chlorinated solvents of concern and their daughter 
products and for total BTEX if present. For example, if trichloroethene and BTEX were released (as 
is typical for fire training areas), then maps of dissolved trichloroethene, dichloroethene, vinyl 
chloride, ethene, and total BTEX concentrations should be prepared. Isopleth maps allow interpreta¬ 
tion of data on the distribution and the relative transport and degradation rates of contaminants in the 
subsurface. In addition, contaminant isopleth maps allow contaminant concentrations to be gridded 
and used for input into a solute transport model. 

Isopleth maps are prepared by first plotting the concentration of the contaminant on a base map 
prepared using surveyor’s data. Lines of equal contaminant concentration (isopleths) are then drawn 


C2-14 


















and labeled. It is important to ensure that each data point is honored during contouring. Outliers 
should be displayed and qualified, if they are not contoured. Figures C.2.4, C.2.7, and C.2.8 contain 
examples of contaminant isopleth maps. 

Dissolved contaminant concentrations are determined through ground-water sampling and 
laboratory analysis. From these data, isopleth maps for each of the contaminant compounds and for 
total dissolved contaminant should be made. Dissolved BTEX concentrations are transferred to the 
fate and transport model grid cells by overlaying the isopleth map onto the model grid. 

C.2.3.6 Preparation of Electron Donor, Inorganic Electron Acceptor, and Metabolic By¬ 
product Contour (Isopleth) Maps 

Isopleth maps should be prepared for any organic compound that can be used as an electron 
donor. Examples of such compounds include natural organic carbon, and petroleum hydrocarbons 
(and landfill leachate). These maps are used to provide visible evidence that biodegradation could 
occur or is occurring. Isopleth maps also should be prepared for dissolved oxygen, nitrate, 
manganese (II), iron (II), sulfate, methane, and chloride. These maps are used to provide visible 
evidence that biodegradation is occurring. The electron acceptor and metabolic by-product isopleth 
maps can be used to determine the relative importance of each of the terminal electron-accepting 
processes (TEAPs). 

Isopleth maps are prepared by first plotting the concentration of the electron donor, electron 
acceptor, or metabolic by-product on a base map prepared using surveyor’s data. Lines of equal 
concentration (isopleths) are then drawn and labeled. It is important to ensure that each data point is 
honored during contouring, unless some data are suspect. 

C.2.3.6.1 Inorganic Electron Acceptor Isopleth Maps 

Electron acceptor isopleth maps allow interpretation of data on the distribution of dissolved 
oxygen, nitrate, and sulfate in the subsurface. Isopleth maps for these compounds provide a visual 
indication of the relationship between the contaminant plume and the electron acceptors and the 


TOTAL BTEX (^g/L) 


TRICHLOROETHENE (mg/L) DICHLOROETHENE (mg/L) 




SULFATE (mg/L) 



FEET 


Figure C.2.7 Example isopleth maps of contaminants and soluble electron acceptors. 


C2-15 











relative importance of each TEAR Dissolved oxygen concentrations below background levels in 
areas with high organic carbon concentrations are indicative of aerobic respiration. Nitrate concen¬ 
trations below background in areas with high organic carbon concentrations are indicative of denitri¬ 
fication. Sulfate concentrations below background in areas with high organic carbon concentrations 
are indicative of sulfate reduction. 

Figure C.2.7 gives examples of completed isopleth maps for dissolved oxygen, nitrate, and 
sulfate. This figure also contains isopleth maps for TCE and DCE and the total BTEX (electron 
donor) isopleth map for the same period. Comparison of the total BTEX isopleth map and the 
electron acceptor isopleth maps shows that there is a strong correlation between areas with elevated 
organic carbon and depleted electron acceptor concentrations. The strong correlation indicates that 
the electron acceptor demand exerted during the metabolism of BTEX has resulted in the depletion 
of soluble inorganic electron acceptors. These relationships provide strong evidence that biodegra¬ 
dation is occurring via the processes of aerobic respiration, denitrification, and sulfate reduction. 



Figure C. 2.8 Example isopleth maps of contaminants and metabolic by-products. 


C2-16 










C.2.3.6.2 Metabolic By-product Isopleth Maps 

Metabolic by-product maps should be prepared for manganese (II), iron (II), methane, and 
chloride. The manganese (II) map is prepared in lieu of an electron acceptor isopleth map for 
manganese (IV) because the amount of bioavailable amorphous or poorly crystalline manganese (IV) 
in an aquifer matrix is extremely hard to quantify. The iron (II) map is prepared in lieu of an electron 
acceptor isopleth map for iron (IE) because the amount of bioavailable amorphous or poorly crystal¬ 
line iron (III) in an aquifer matrix is extremely hard to quantify. Iron (E) concentrations above 
background levels in areas with BTEX contamination are indicative of anaerobic iron (IE) reduction. 
Methane concentrations above background levels in areas with BTEX contamination are indicative 
of methanogenesis, another anaerobic process. Biodegradation of chlorinated solvents tends to 
increase the chloride concentration found in ground water. Thus, chloride concentrations inside the 
contaminant plume generally increase to concentrations above background. This map will allow 
visual interpretation of chloride data by showing the relationship between the contaminant plume 
and chloride. During anaerobic biodegradation, the oxidation-reduction potential of ground water is 
lowered. Thus, the oxidation-reduction potential (or pE) inside the contaminant plume generally 
decreases to levels below background. 

Figure C.2.8 gives examples of completed isopleth maps for iron (II), methane, chloride, and 
pE. This figure also contains the TCE, DCE and Vinyl Chloride isopleth maps, and total BTEX 
(electron donor) isopleth map for the same period. Comparison of the total BTEX isopleth map and 
the metabolic by-product isopleth maps and comparison of Figures C.2.7 and C.2.8 shows that there 
is a strong correlation between areas with elevated organic carbon and elevated metabolic by-product 
concentrations. These relationships provide strong evidence that biodegradation is occurring via the 
processes of iron (IE) reduction, methanogenesis, and reductive dechlorination. 


C2-17 


SECTION C-3 

NATURAL ATTENUATION CALCULATIONS 

Several calculations using site-specific data must be made in order to document the occurrence 
of natural attenuation and successfully implement the natural attenuation alternative. The following 
sections describe these calculations. 

C.3.1 CALCULATING HYDRAULIC PARAMETERS 

Hydraulic parameters necessary for adequate site characterization and model implementation 
include hydraulic conductivity, transmissivity, hydraulic gradient, linear ground-water flow velocity, 
hydrodynamic dispersion, and retarded solute transport velocity. Calculations for these parameters 
are discussed in the following sections. 

C.3.1.1 Hydraulic Conductivity 

Hydraulic conductivity, K, is a measure of an aquifer’s ability to transmit water and is perhaps 
the most important variable governing fluid flow in the subsurface. Hydraulic conductivity has the 
units of length over time [L/T]. Observed values of hydraulic conductivity range over 12 orders of 
magnitude, from 3x10 12 to 3 cm/sec (3x1 O' 9 to 3x10 3 m/day) (Figure C.3.1 and Table C.3.1). In 


Unconsolidated 

Deposits 

Gravel 

coats* 

mecfium 

fin* 

Sand 

coats* 

madum 

fin* 

Alluvial deposits 

Silt 

Clay 

d*nse 

wealhered 

Rocks 

Sandstone 

dense 

katstfc 

Limestone 

dense 

karstic 

Dolomite 
Crystafline rocks 

danse 

fractured 

Basalt 

d*ne* 

fractured 

Claystone 

Volcanic tuff 

Shale 

dens* 

fractured 


■■ typical range 

Modified from: Spitz and Mortno, 1996. 

Figure C.3. 1 Range of hydraulic conductivity values. 


K in cm/s 

practically 

low 


high 

impermeable 

permeability 

permeable 

permeability 

1 C 9 

10 7 10" 5 

10" 3 10' 1 10 



C3-18 











































general terms, the hydraulic conductivity for unconsolidated sediments tends to increase with in¬ 
creasing grain size and sorting. The velocity of ground water and dissolved contaminants is directly 
related to the hydraulic conductivity of the saturated zone. Subsurface variations in hydraulic con¬ 
ductivity directly influence contaminant fate and transport by providing preferential pathways for 
contaminant migration. The most common methods used to quantify hydraulic conductivity in the 
subsurface are aquifer pumping tests and slug tests. The quantitative analysis of pumping and slug 
test data is beyond the scope of this document. For information on the quantitative analysis of these 
data, the reader is referred to the works of Kruseman and de Ridder (1991) and Dawson and Istok 
(-1991). 


Table C.3.1 Representative Values of Hydraulic Conductivity for Various Sediments and Rocks (From 
Domenico and Schwartz, 1990) 


Material 

Hydraulic Conductivity 
(m/day) 

Hydraulic Conductivity 
(cm/sec) 

UNCONSOLIDATED 

SEDIMENT 



Glacial till 

9x1 O' 8 - 2x1 O' 1 

lxlO' 10 - 2x10^ 

Clay 

9x1 O' 7 - 4x1 O' 4 

lxlO' 9 - 5xl0' 7 

Silt 

9x10‘ 5 - 2 

lxlO' 7 - 2xl0' 3 

Fine sand 

2xl0' 2 -2xl0 1 

2xl0' 5 - 2xl0' 2 

Medium sand 

8x1 O' 2 - 5x10* 

9x1 O' 5 - 6x1 O' 2 

Coarse sand 

8x1 O' 2 - 5xl0 2 

9x1 O' 5 - 6x1 O' 1 

Gravel 

3x10' - 3xl0 3 

3xl0' 2 - 3 

SEDIMENTARY ROCK 



Karstic limestone 

9x1 O' 2 - 2x10 3 

lxlO” 4 -2 

Limestone and dolomite 

9x1 O' 5 - 5x1 O' 1 

lxlO' 7 - 6x10" 4 

Sandstone 

3xl0' 5 - 5X10' 1 

3x1 O' 8 - 6x10^ 

Siltstone 

9x1 O' 7 - lxl O' 3 

lxlO' 9 - lxlO' 6 

Shale 

9xl0' 9 - 2x1 O' 4 

lxlO' 11 - 2xl0' 7 

CRYSTALLINE ROCK 



Vesicular basalt 

3xl0' 2 - 2xl0 3 

4x1 O' 5 - 2 

Basalt 

2x1 O’ 6 - 3x1 O' 2 

2xl0' 9 - 4xl0' 5 

Fractured igneous and 

metamorphic 

7x1c 4 - 3x10’ 

8xl0' 7 - 3xl0' 2 

Unfractured igneous 
and metamorphic 

3xl0' 9 - 2x1 O' 5 

3xl0' 12 - 2xl0' 8 


C3-19 

























C.3.1.1.1 Hydraulic Conductivity from Pumping Tests 

Pumping tests generally provide the most reliable information about aquifer hydraulic conduc¬ 
tivity. Pumping test data for geohydraulic characteristics are most commonly interpreted by graphic 
techniques. The analytical method used for interpretation of the data will depend upon the physical 
characteristics of the aquifer and test wells. The assumptions inherent in the analytical method used 
to calculate aquifer characteristics should be evaluated to ensure acceptance of the method for the 
subsurface conditions present at the site under investigation. 

The interpretation of aquifer pumping test data is not unique. Similar sets of data can be ob¬ 
tained from various combinations of geologic conditions. Field data of drawdown vs. time and/or 
distance are plotted on graph paper either by hand or using programs such as AQTESOLV® or a 
spreadsheet program. There are numerous methods of interpreting pumping test data. The method 
to be used for each pumping test should be selected based on site-specific conditions (aquifer condi¬ 
tions, test conditions, assumptions made, etc.). Most hydrogeology text books contain pumping test 
evaluation techniques. Two publications dealing with pump test analysis are recommended 
(Kruseman and de Ridder, 1991 and Dawson and Istok, 1991). 

C.3.1.1.2 Hydraulic Conductivity from Slug Tests 

Slug tests are a commonly used alternative to pumping tests that are relatively easy to conduct. 
The biggest advantage of slug tests is that no contaminated water is produced during the test. During 
pumping tests at fuel-hydrocarbon-contaminated sites, large volumes of contaminated water that 
must be treated typically are produced. One commonly cited drawback to slug testing is that this 
method generally gives hydraulic conductivity information only for the area immediately surround¬ 
ing the monitoring well. If slug tests are going to be relied upon to provide information on the three- 
dimensional distribution of hydraulic conductivity in an aquifer, multiple slug tests must be per¬ 
formed, both within the same well and at several monitoring wells at the site. It is not advisable to 
rely on data from one slug test in a single monitoring well. Data obtained during slug testing are 
generally analyzed using the method of Hvorslev (1951) for confined aquifers or the method of 
Bouwer and Rice (1976) and Bouwer (1989) for unconfined conditions. 

C.3.1.2 Transmissivity 

The transmissivity, T, of an aquifer is the product of the aquifer’s hydraulic conductivity, K, and 
the saturated thickness, b: 

T=Kb eq. C.3.i 

For a confined aquifer, b is the thickness of the aquifer between confining units. For uncon- 
fined aquifers, b is the saturated thickness of the aquifer measured from the water table to the under¬ 
lying confining layer. Transmissivity has the units of length squared over time [L 2 /T]. 

C.3.1.3 Hydraulic Head and Gradient 

Determining the magnitude of hydraulic gradients is important because gradients influence the 
direction and rate of contaminant migration. Hydraulic head, H, and specifically, variations in 
hydraulic head within an aquifer, is the driving force behind ground-water movement and solute 
migration. The total hydraulic head at one location in a system is the sum of the elevation head, 
pressure head, and velocity head (Figure C.3.2): 

H=h z + h p +h v eq. C.3.2 

Where: 

H = total hydraulic head [L] 

h z = elevation head = z = elevation relative to the reference plane [L] 
h p = pressure head [L] 
h v = velocity head [L] 


C3-20 


Pressure head is given by: 


Where: 

p = fluid pressure 
p = density 

g = acceleration due to gravity 
Velocity head is given by: 


Where: 

v = ground-water velocity 
g = acceleration due to gravity 



P_ 

PS 



Because h v is generally assumed to be zero for most ground-water flow, the relationship for total 
head is generally written: 


H=z 



eq. C.3.3 


Thus, the total hydraulic head at a point measured by a piezometer is the sum of the elevation at the 
base of the piezometer plus the length of the water column in the piezometer. The total hydraulic 
head in a piezometer is determined by measuring the depth from a surveyed reference point (datum) 
to the surface of the standing water. The elevation of the water surface is the total hydraulic head in 
the piezometer. This total head is the total head at the base of the piezometer, not the water table 
elevation, unless the piezometer terminates immediately below the water table or is a well screened 


A 


Depth to Water 

} t 


Pressure 
Head p 


B /-Measurement Datum 

Open-ended Tube (Piezometer) 




Depth to Water 


T 


Hea ^ 


-Pressure 


Ground Surface 


y Water Tabl e 


Total Head (H) 


Total Head (t ) 


Elevation 
Head (z) 


Elevation 
Head (z) 


1 i. 


i _±- 

Mean Sea Level (Reference Elevation 


Figure C.3.2 Hydraulic head. 


C3-21 
























across the water table. Figure C.3.2 shows a pair of nested piezometers that illustrate the relation¬ 
ships between total hydraulic head, pressure head, and elevation head. Because ground water flows 
from areas with high total head (point A, Figure C.3.2) to areas with lower total head (point B), this 
figure depicts a water table aquifer with a strong upward vertical gradient. This figure illustrates 
how nested piezometers (or wells) are used to determine the importance of vertical gradients at a 
site. This figure also illustrates the importance of using wells screened in the same portion of the 
aquifer (preferably across the water table) when preparing potentiometric surface maps. 

The hydraulic gradient (dH/dL) is a dimensionless number that is the change in hydraulic head 
(dH) between two points divided by the length of ground-water flow between these same two points, 
parallel to the direction of ground-water flow, and is given by: 


Hydraulic Gradient = 


dH 

dL 


eq. C.3.4 


Where: 

dH = change in total hydraulic head between two points [L] 

dL = distance between the two points used for head measurement [L] 

In a system where flow is not occurring, the total hydraulic head, H, is the same everywhere in 
the system and the hydraulic gradient is zero. To accurately determine the hydraulic gradient, it is 
necessary to measure ground-water levels in all monitoring wells at the site. Because hydraulic 
gradients can change over a short distance within an aquifer, it is essential to have as much site- 
specific ground-water elevation information as possible so that accurate hydraulic gradient calcula¬ 
tions can be made. In addition, seasonal variations in ground-water flow direction can have a pro¬ 
found influence on contaminant transport. To determine the effect of seasonal variations in ground- 
water flow direction on contaminant transport, quarterly ground-water level measurements should be 
taken over a period of at least 1 year. 

The hydraulic gradient must be determined parallel to the direction of ground-water flow. 

Unless two monitoring wells screened in the same relative location within the same hydrogeologic 
unit are located along a line parallel to the direction of ground-water flow, the potentiometric surface 
map is generally used to determine the hydraulic gradient. To determine the hydraulic gradient, an 
engineer’s scale is used to draw a line perpendicular to the equal-potential lines on the potentiomet¬ 
ric surface map (i.e., parallel to the direction of ground-water flow). Measure the distance between 
the two equal-potential lines, making note of the ground-water potential at each equal-potential line. 
Subtract the larger potential from the smaller potential, and divide this number by the distance 
between the two equal potential lines, being sure to use consistent units. The number generated will 
be a negative number because water flows from areas of higher potential to areas of lower potential. 

Example C.3.1 : Hydraulic Gradient Calculation 

Given the water table elevation map shown in Figure C.3.3, calculate the hydraulic gradient 
between points A and B. Assume that all wells are screened across the water table. 

Solution: 

The hydraulic gradient is given by dH/dL. The line connecting points A and B is parallel to the 
direction of ground-water flow. The water table elevation is 4659.34 ft msl at point A and 
4602.41 ft msl at point B. Therefore, because ground water flows from areas of high head to areas of 
lower head: 

dH- 4602.41 - 4659.34 = - 56.93 feet 
The distance between the two points A and B is 936 feet. Therefore: 

dL- 936 feet 


C3-22 






Figure C.3.3 Ground water elevation map. 


and 


dH -56.93 ft 
dL~ 93 6 ft 


9.06—=-0.06 — 


ft 


m 


C.3.1.4 Total Porosity (n) and Effective Porosity (n e ) 

Total porosity (n) is the volume of voids in a unit volume of aquifer. Specific retention is the 
amount of water (volumetric) that is retained against the force of gravity after a unit volume of an 
unconfined aquifer is drained. Storativity is defined as the volume of water that a confined aquifer 
takes into or releases from storage per unit surface area of the aquifer per unit change in total hydrau¬ 
lic head. Effective porosity, n e , is the total porosity of the aquifer minus the specific retention (un¬ 
confined) or storativity (confined) of the aquifer: 

n e =n-S eq. C.3.5 

Where: 

n e - effective porosity [dimensionless] 
n = total porosity [dimensionless] 

S = specific retention (unconfined) or storativity (confined) [dimensionless] 

Effective porosity can be estimated using the results of a tracer test. Although this is potentially the 
most accurate method, time and monetary constraints can be prohibitive. For this reason, the most 
common technique is to use an accepted literature value for the types of materials making up the 
aquifer matrix, and then to calibrate a contaminant transport model by adjusting the value of effec¬ 
tive porosity (in conjunction with other input parameters such as transmissivity) within the range of 


C3-23 








accepted literature values until the modeled and observed contaminant distribution patterns match. 
Because aquifer materials can have a range of effective porosity, sensitivity analyses should be 
performed to determine the effect of varying the effective porosity on numerical model results. 
Values of effective porosity chosen for the sensitivity analyses should vary over the accepted range 
for the aquifer matrix material. Table C.3.2 presents accepted literature values for total porosity and 
effective porosity. 

Table C.3.2 Representative Values of Dry Bulk Density, Total Porosity, and Effective Porosity for 

Common Aquifer Matrix Materials (After Walton, 1988 and Domenico and Schwartz, 1990) 


Aquifer 

Matrix 

Dry Bulk 

Density 

(gm/cm 3 ) 

Total 

Porosity 

Effective 

Porosity 

Clay 

1.00-2.40 

0.34- 

0.60 

0.01-0.2 

Peat 

— 

— 

0.3-0.5 

Glacial 

Sediments 

1.15-2.10 

— 

0.05-0.2 

Sandy Clay 

— 

— 

0.03-0.2 

Silt 

— 

0.34- 

0.61 

0.01-0.3 

Loess 

0.75-1.60 

— 

0.15-0.35 

Fine Sand 

1.37-1.81 

0.26- 

0.53 

0.1-0.3 

Medium Sand 

1.37-1.81 

— 

0.15-0.3 

Coarse Sand 

1.37-1.81 

0.31- 

0.46 

0.2-0.35 

Gravely Sand 

1.37-1.81 

— 

0.2-0.35 

Fine Gravel 

1.36-2.19 

0.25- 

0.38 

0.2-0.35 

Medium 

Gravel 

1.36-2.19 

— 

0.15-0.25 

Coarse Gravel 

1.36-2.19 

0.24- 

0.36 

0.1-0.25 

Sandstone 

1.60-2.68 

0.05- 

0.30 

0.1-0.4 

Siltstone 

— 

0.21- 

0.41 

0.01-0.35 

Shale 

1.54-3.17 

0.0-0.10 

— 

Limestone 

1.74-2.79 

0.0-50 

0.01-0.24 

Granite 

2.24-2.46 

— 

— 

Basalt 

2.00-2.70 

0.03- 

0.35 

— 

Volcanic Tuff 

— 

— 

0.02-0.35 


C.3.1.5 Linear Ground-water Flow Velocity (Seepage or Advective Velocity) 

The average linear ground-water flow velocity (seepage velocity) in one dimension in the 
direction parallel to ground-water flow in a saturated porous medium is given by: 


C3-24 


























K dH 
n e dL 


eq. C.3.6 


v 


x 


average linear ground-water velocity parallel to ground-water flow direction (seepage 
velocity) [L/T] 
hydraulic conductivity [L/T] 
effective porosity [L 3 /L 3 ] 

= hydraulic gradient [L/L] 

The average linear ground-water flow velocity should be calculated to estimate ground-water flow 
and solute transport velocity, to check the accuracy of ground-water models, and to calculate first- 
order biodegradation rate constants. 

Example C.3.2 : Linear Ground-water Flow Velocity Calculation 

Calculate the linear ground-water flow velocity in a medium-grained sandy aquifer. The hy¬ 
draulic gradient as determined from the potentiometric surface map in the previous example is - 
0.06 m/m. The nydraulic conductivity is 1.7x10'’ m/day as determined by pumping tests. 

Solution: 

Because the effective porosity of this sediment is not known, it is necessary to estimate this 
parameter. From Table C.3.2, the effective porosity for a medium-grained sand is approximately 
23 percent. 


Where: 

v = 
* 

K = 
n = 

e 

dH 

dL 


K dH 
n e dL 


( 0A7 %ay)(-°- 06m /J 

0.23 



C.3.1.6 Coefficient of Retardation and Retarded Contaminant Transport Velocity 

When the average linear velocity of a dissolved contaminant is less than the average linear 
velocity of the ground water, the contaminant is said to be “retarded.” The difference between the 
velocity of the ground water and that of the contaminant is caused by sorption and is described by the 
coefficient of retardation, R, which is defined as: 

R = V eq.C.3.7 

v c 

Where: 

R = coefficient of retardation 

v = average linear ground-water velocity parallel to ground-water flow 
v = average velocity of contaminant parallel to groundwater flow 
The ratio v /v describes the relative velocity between the ground water and the dissolved contami¬ 
nant. When K d = 0 (no sorption), the transport velocities of the ground water and the solute are equal 
(v = v ). If it can be assumed that sorption is adequately described by the distribution coefficient, 
the coefficient of retardation for a dissolved contaminant (for saturated flow) is given by: 

= l + eq.C.3.8 

n 

Where: 

R = coefficient of retardation 
p b = bulk density (Section C.3.1.6.1) 

K = distribution coefficient (Section C.3.1.6.2) 
n = total porosity 


C3-25 









This relationship expresses the coefficient of retardation in terms of the bulk density and effective 
porosity of the aquifer matrix and the distribution coefficient for the contaminant. Substitution of 
this equation into equation C.3.7 gives: 



A** 

n 


eq. C.3.9 


Solving for the contaminant velocity, v c , gives: 


Vc i | A k < eq.C.3.10 

n 

Retardation of a contaminant relative to the advective transport velocity of the ground-water flow' 
system has important implications for natural attenuation. If retardation is occurring, dissolved 
oxygen and other electron acceptors traveling at the advective transport velocity of the ground water 
sweep over the contaminant plume from the upgradient margin. This results in greater availability of 
electron acceptors within the plume for biodegradation of fuel hydrocarbons. In addition, adsorption 
of a contaminant to the aquifer matrix results in dilution of the dissolved contaminant plume. 

C.3.1.6.1 Bulk Density 

The bulk density of a soil, p b , as used in most ground-water models, expresses the ratio of the 
mass of dried soil to its total volume (solids and pores together). 

. K K 

V T (K + K + K) eq.c.3.11 


Where: 

p h - bulk density 

M= mass of solid in the system 

V T = total volume in the system 

V = volume of solid in the system 

V = volume of air (or gas) in the system 

F = volume of water (or liquid) in the system 
Bulk density is related to particle density by: 

A =.(!-»)/? eq. C.3.H 

Where: 

p b = bulk density 

n = total porosity 

p t = density of grains comprising the aquifer 

The bulk density is always less than the particle density, p s ; for example, if pores constitute half 
the volume, then p b is half of p s . The bulk density of a soil is affected by the structure of the soil 
(looseness and degree of compaction), as well as by its swelling and shrinking characteristics, both 
of which depend on clay content and soil moisture. Even in extremely compacted soil, the bulk 
density remains appreciably lower than the particle density. This is because the particles can never 
interlock perfectly, and the soil remains a porous body, never completely impervious. In sandy soils, 
p b can be as high as 1.81 gm/cm 3 . In aggregated loams and clayey soils, p b can be as low as 
1.1 gm/cm 3 . Table C.3.2 contains representative values of dry bulk density for common sediments 
and rocks. 


C.3.1.6.2 Distribution Coefficient and Total Organic Carbon Content 

The distribution coefficient is described in Section B.4.3. Recall equation B.4.10, which gives 
the relationship between f and K : 

r J oc oc 


C3-26 







K, =Kf 

a ocJ oc 


eq. C.3.13 


Where: 

K d = distribution coefficient [L 3 /M] 

K o = soil adsorption coefficient for soil organic carbon content [L 3 /M] 
f oc = fraction soil organic carbon (mg organic carbon/mg soil) [M/M] 

Representative K oc values are given in Table B.4.1. The fraction of soil organic carbon must be 
determined from site-specific data. Representative values of total organic carbon (TOC) in common 
sediments are given in Table C.3.3. Because most solute transport occurs in the most transmissive 
aquifer zones, it is imperative that soil samples collected for total organic carbon analyses come from 
these zones in background areas. To be conservative, the average of all total organic carbon concen¬ 
trations from sediments in the most transmissive aquifer zone should be used for retardation calcula¬ 
tions. 

Table C.3.3 Representative Values of Total Organic Carbon for Common Sediments 


Texture 

Depositional Environment 

Fraction Organic 

Carbon 

Site Name 

medium sand 

fluvial-deltaic 

0.00053 - 0.0012 

Hill AFB, Utah 

fine sand 


0.0006 - 0.0015 

Bolling AFB, D.C. 

fine to coarse sand 

back-barrier (marine) 

0.00026 - 0.007 

Patrick AFB, Florida 

organic silt and peat 

glacial (lacustrine) 

0.10 - 0.25 

Elmendorf AFB, Alaska 

silty sand 

glacio fluvial 

0.0007 - 0.008 

Elmendorf AFB, Alaska 

silt with sand, gravel and 
clay (glacial till) 

glacial moraine 

0.0017 - 0.0019 

Elmendorf AFB, Alaska 

medium sand to gravel 

glaciofluvial 

0.00125 

Elmendorf AFB, Alaska 

loess (silt) 

eolian 

0.00058 - 0.0016 

Offutt AFB, Nebraska 

fine - medium sand 

glaciofluvial or 
glaciolacustrine 

< 0.0006 - 0.0061 

Truax Field, Madison 

W isconsin 

fine to medium sand 

glaciofluvial 

0.00021 - 0.019 

King Salmon AFB, Fire 

Training Area, Alaska 




Dover AFB, Delaware 

fine to coarse sand 

glaciofluvial 

0.00029 - 0.073 

Battle Creek ANGB, Michigan 

sand 

fluvial 

0.0057 

Oconee River, Georgia 47 

coarse silt 

fluvial 

0.029 

Oconee River, Georgia 47 

medium silt 

fluvial 

0.020 

Oconee River, Georgia 47 

fine silt 

fluvial 

0.0226 

Oconee River, Georgia 47 

silt 

lacustrine 

0.0011 

Wildwood, Ontario b 

fine sand 

glaciofluvial 

0.00023 - 0.0012 

Various sites in Ontario 1 " 7 

medium sand to gravel 

glaciofluvial 

0.00017 - 0.00065 

Various sites in Ontario b/ 


a/ Karickhoff, 1981 

b/ Domenico and Schwartz (1990) 


Example C.3.3 : Retarded Solute Transport Velocity Calculation 

For ground-water flow and solute transport occurring in a shallow, saturated, well-sorted, fine¬ 
grained, sandy aquifer, with a total organic carbon content of 0.7 percent, a hydraulic gradient of - 
0.015 m/m, and an hydraulic conductivity of 25 m/day, calculate the retarded contaminant velocity 
for trichloroethene. 

Solution: 

Because the total porosity, effective porosity, and the bulk density are not given, values of 
these parameters are obtained from Table C.3.2. The median values for total porosity, effective 


C3-27 





























porosity, and bulk density are approximately 0.4,0.2, and 1.6 kg/L, respectively. 
The first step is to calculate the average linear ground-water velocity, v x . 


v 


x 



The next step is to determine the distribution coefficient, K d . Values of K oc for chlorinated 
solvents and BTEX are obtained from Tables B.2.1 and B.2.2, respectively, and are listed in 
Table C.3.4. 

For trichloroethene = 87 L/kg, and (using equation C.3.13): 



( j \ L 

87— (0.007) = 0.61 — 

V k SJ k g 


The retarded contaminant velocity is given by (equation C.3.10): 


v 


c 



Table C.3.4 presents the estimated coefficient of retardation contaminant velocity for a number of 
contaminants under the conditions of Example C.3.3. This example illustrates that contaminant 
sorption to total organic carbon can have a profound influence on contaminant transport by signifi¬ 
cantly slowing the rate of dissolved contaminant migration. 


Table C.3.4 Example Retardation Calculations for Select Compounds 


Gonpound 

Koc 

L/kg 

Fraction 

Organic 

Carbon 

Distillation 

Coefficient 

(L/kg) 

Bdk 

Density 

(kg/L) 

Total 

Rwoshy 

Coefficient of 

Retardation 

Advective 

Gnxmd-vvater 
Velocity' (nfday) 

Cbntaninant 

Velocity' 

(rrtday) 

Benzene 

79 

0.007 

0.553 

1.60 

0.40 

3.21 

1.90 

0.59 

Toluene 

190 

0.007 

1.33 

1.60 

0.40 

6.32 

1.90 

0.30 

Ethylbenzene 

468 

0.007 

3.276 

1.60 

0.40 

14.10 

1.90 

0.13 

m-xylene 

405 

0.007 

2.835 

1.60 

0.40 

12.34 

1.90 

0.15 

Tetrachbroethene 

209 

0.007 

1.463 

1.60 

0.40 

6.85 

1.90 

0.28 

Tridibroethene 

87 

0.007 

0.609 

1.60 

0.40 

3.44 

1.90 

0.55 

cis-1,2-Dichloroetbene 

49 

0.007 

0.343 

1.60 

0.40 

2.37 

1.90 

0.80 

Vinyl Chloride 

2.5 

0.007 

0.0175 

1.60 

0.40 

1.07 

1.90 

1.78 

1 3,5-tiiiHhyiben2ene 

676 

0.007 

4.732 

1.60 

0.40 

19.93 

1.90 

0.10 


C.3.2 CONTAMINANT SOURCE TERM CALCULATIONS 

NAPLs present in the subsurface represent a continuing source of ground-water contamination. 
NAPLs may be made up of one compound, or more likely, a mixture of compounds. Concentrations 
of dissolved contaminants and the lifetime of NAPL source areas and associated ground-water 
plumes are ultimately determined by the rate at which contaminants dissolve from the NAPL. When 
sufficient quantities of NAPL are present, the unsaturated zone may initially be saturated with 


C3-28 

















NAPL, and the NAPL may migrate under the influence of gravity. After a period of time the NAPL 
may drain from the pores under the influence of gravity, leaving a thin coating of NAPL. Depending 
on the surface area of the subsurface materials, the surface tension of the NAPL, and the porosity and 
permeability of the subsurface materials, some NAPL also may be held between the grains by capil¬ 
larity. NAPL adhering to the grains of the aquifer matrix or retained by capillarity is herein referred 
to as residual NAPL. In residual zones, NAPL will be present in immobile blobs or ganglia that may 
occupy 10 percent or less of the pore space (Feenstra and Guiguer, 1996). If the NAPL is at satura¬ 
tion and is mobile within and among the pores of the aquifer matrix, the NAPL is referred to as 
mobile NAPL. Mobile NAPL may occupy as much as 50 to 70 percent of the pore space and can 
reduce flow of water through these zones. 

In the unsaturated zone, dissolution from residual or mobile NAPL into downward-migrating 
precipitation (recharge) will occur, as well as migration and dissolution of vapors. In the saturated 
zone, dissolution of contaminants from residual NAPL occurs as ground-water flows through the 
residual zone. Dissolution from mobile NAPL mostly takes place along the tops, bottoms, or lateral 
margins of the NAPL bodies, because ground-water (or recharge) flow through the NAPL is re¬ 
stricted. Because the distribution of residual NAPL results in a greater surface area of product in 
contact with ground water and does not restrict ground-water velocities, concentrations of contami¬ 
nants entering ground water will typically be closer to the compounds’ equilibrium solubilities than 
in the case of mobile NAPL bodies. The equilibrium solubility of the compound(s) of interest will 
depend on the composition of the NAPL (i.e., the molar fraction of the NAPL represented by the 
compound). 

In general, residual and mobile NAPL may be present above or below the water table, but direct 
dissolution into ground water will only occur when NAPL is at or below the capillary fringe. In 
either case, quantifying the flux of contamination entering ground water from above or below the 
water table is a difficult proposition. The processes governing dissolution from NAPLs are complex 
and depend upon many variables (Feenstra and Guiguer, 1996). Among these variables (in the 
saturated zone) are the shape of a mobile NAPL body, the contact area between the NAPL and the 
ground water, the velocity of the ground water moving through or past the NAPL, the effect of 
residual NAPL on the effective porosity of the contact zone, the solubility of the compounds of 
interest, the relative fractions of the compounds in the NAPL, the diffusion coefficients of the com¬ 
pounds, and the effects of other compounds present in the NAPL. This will be further complicated 
by any processes in the vadose zone (e.g., volatilization, dissolution from residual NAPL into re¬ 
charge, or dissolution of vapors into recharge) that also will add contaminant mass to ground water. 
Further, as the mass of the NAPL body changes over time, the rate of dissolution will also change. 
Clearly, given the number of variables that affect the transfer of contaminant mass to ground water, it 
is difficult to accurately estimate the flux of contaminants into ground water. Depending on the 
intended use of the flux estimate, different approaches can be used. 

If one desires to estimate a source term for a contaminant fate and transport model, one can 
attempt to estimate the mass loading rate and use that estimate as an input parameter. However, this 
often does not yield model concentrations (dissolved) that are similar to observed concentrations. As 
a result, the source in the model often becomes a calibration parameter (Mercer and Cohen, 1990; 
Spitz and Moreno, 1996). This is because the effects of the source (i.e., the dissolved contaminant 
plume) are easier to quantify than the actual flux from the source. The frequent need for such a 
“black box” source term has been borne out during modeling associated with evaluations of natural 
attenuation of fuel hydrocarbons [following the AFCEE technical protocol (Wiedemeier et 
fl/.,1995d)] at over 30 U.S. Air Force sites. Use of other methods to calculate source loading for 
those models often produced model concentrations that differed from observed concentrations by as 


C3-29 


much as an order of magnitude. From the model, the flux estimate then can be used for estimating 
source lifetimes or other such calculations. 

For other purposes, one can estimate flux using several methods, as summarized by Feenstra 
and Guiguer (1996). For bodies of mobile LNAPL, this is more practical, because the area of NAPL 
in contact with ground water can be estimated from plume/pool dimensions. Where most NAPL is 
residual, the surface area can be highly variable, and cannot be measured in the field. Laboratory 
studies to understand and quantify mass transfer from residual NAPL in porous media are in the 
early stages, and when such mass transfer is modeled, surface area is a calibration parameter with 
great uncertainty (Abriola, 1996). Most methods of estimating NAPL dissolution rates require an 
estimate of the contact area and, therefore, will contain a great deal of uncertainty. This is one of the 
main reasons why, for purposes of modeling, the “black box” source term is more commonly used. 

One reason practitioners want to estimate mass transfer rates is to provide a basis for estimating 
contaminant source lifetimes, which can affect regulatory decisions and remedial designs. To deter¬ 
mine how long it will take for a dissolved contaminant plume to fully attenuate, it is necessary to 
estimate how fast the contaminants are being removed from the NAPL. In general, it is difficult to 
estimate cleanup times, so conservative estimates should be made based on NAPL dissolution rates. 
Predicting the cleanup time for sites with mobile NAPL is especially difficult because residual 
NAPL will remain after the recoverable mobile NAPL has been removed. Of course, this is all 
complicated by the many factors that affect dissolution rates as discussed above. Moreover, most 
methods do not account for changing dissolution rates as a result of NAPL volume loss (and subse¬ 
quent surface area decrease), preferential partitioning from mixed NAPLs, and the change in porosity 
(and, therefore, ground-water velocity) resulting from NAPL dissolution. Finally, the mass of the 
NAPL present in the subsurface must also be estimated, lending further uncertainty to any calcula¬ 
tion of source lifetime. 

There are several ways to quantify the mass loading rate from a body of mobile or residual 
NAPL. Feenstra and Guiguer (1996) present a good summary of some common methods. As noted 
above, transfer rates calculated from these methods are all dependent upon several parameters, many 
of which cannot be measured or derived from the literature. This is especially true for residual 
NAPL. Johnson and Pankow (1992) present a method for estimating dissolution rates from pools of 
NAPL which contact ground water over an area that is essentially two-dimensional. Many other 
dissolution models may be available; however, as noted before, the experimental evidence to support, 
dissolution models is really just starting to be collected. Despite these limitations, some of these 
models can prove useful, and a selected few are presented (in limited detail) in the following subsec¬ 
tions. 

If estimating mass flux rates is less important, one can use direct measurement or equilibrium 
concentration calculations to estimate contaminant source area concentrations. The first method 
involves directly measuring the concentration of dissolved contaminants in ground water near the 
NAPL plume. The second method involves the use of partitioning calculations. These approaches 
are described in the following sections. This type of data can be useful if it can be demonstrated that 
the source is not capable of introducing concentrations of compounds of concern that exceed regula¬ 
tory limits, or that with slight weathering the same results can be expected. Source area concentra¬ 
tions, whether measured or calculated, also may be used to provide calibration targets for transport 
models in which a “black box” source term is used. 

If contaminant concentrations in the residual and mobile NAPL are not decreasing over time, or 
if they are decreasing very slowly, extremely long times will be required for natural attenuation of the 
dissolved contaminant plume. This will likely make natural attenuation less feasible and will reduce 
the chance of implementation. In order for natural attenuation to be a viable remedial option, the 


C3-30 


source of continuing ground-water contamination must be decreasing over time (decaying), either by 
natural weathering processes or via engineered remedial solutions such as mobile NAPL recovery, 
soil vapor extraction, bioventing, or bioslurping. Because natural weathering processes can be fairly 
slow, especially in systems where the NAPL dissolves slowly or is inhibited from volatilizing or 
biodegrading, it will generally be necessary to implement engineered remedial solutions to remove 
the NAPL or reduce the total mass of residual and dissolved NAPL. 

A discussion of estimating source terms for sites contaminated solely with fuel hydrocarbons is 
presented by Wiedemeier et al. (1995a). In general, estimating dissolution rates of individual com¬ 
pounds from fuels is simpler than estimating rates of dissolution from other NAPL mixtures because 
there is a great deal of experimental evidence regarding partitioning and equilibrium solubilities of 
individual compounds from common fuel mixtures. Methods presented in the following subsections 
can use such data to reduce some of the uncertainty in source term calculations. 

Typical uses of chlorinated solvents (e.g.., degreasing or parts cleaning) and past disposal 
practices that generally mixed different waste solvents or placed many types of waste solvents in 
close proximity have resulted in complex and greatly varying NAPL mixtures being released at sites. 
For mixtures containing other compounds (e.g., either DNAPLs containing multiple chlorinated 
compounds, or fuel LNAPLs containing commingled chlorinated compounds), the equilibrium 
solubility of the individual compounds of interest 'must first be calculated, then that information can 
be used in the common mass transfer rate calculations. Except in the case of pure solvent spills, 
therefore, the estimation of dissolution rates is then further complicated by this need to estimate 
equilibrium solubilities from the mixture. 

Because this work focuses largely on saturated-zone processes, vadose zone dissolution pro¬ 
cesses will not be discussed in any detail. However, this discussion will provide a starting point for 
estimating source terms for ground-water contaminant fate and transport modeling, as well as for 
estimating source and plume lifetimes. As a starting point, two basic methods of estimating or 
measuring equilibrium dissolved contaminant concentrations in the vicinity of NAPL bodies are 
presented. In addition, methods for estimating fluxes summarized by Feenstra and Guiguer (1996) 
and presented by Johnson and Pankow (1992) will be briefly summarized. 

C.3.2.1 Direct Measurement of Dissolved Contaminant Concentrations in Ground Water in 
Contact with NAPL 

Two methods can be used to determine the dissolved concentration of contaminants in ground 
water near a NAPL plume. The first method involves collecting ground-water samples from near a 
NAPL lens in monitoring wells. The second method involves collecting samples of mixed NAPL 
and water from monitoring wells. 

C.3.2.1.1 Collecting Ground-water Samples from Near the NAPL 

This method involves carefully sampling ground water beneath a floating LNAPL lens or near a 
DNAPL lens. One way of collecting a ground-water sample from beneath a lens of floating LNAPL 
or above/adjacent to a DNAPL body involves using a peristaltic pump. For LNAPL, the depth to the 
base of the mobile LNAPL is measured, a length of high-density polyethylene (HDPE) tubing that 
will reach 1 to 2 feet beneath the LNAPL is lowered into the well, and the sample is collected. For 
DNAPL, the tube would be cut to reach 1 to 2 feet above the NAPL. Another useful technique for 
obtaining such samples where the depth to ground water is too deep to allow use of a peristaltic 
pump is to use a Grundfos® pump. If a Grundfos® pump is used to collect a water sample from 
beneath LNAPL, it is imperative that the pump be thoroughly cleaned after each use, and that good 
sampling logic be used (e.g., sample less contaminated wells first). Also, dedicated bladder pumps 
that are being used for long-term monitoring (LTM) in wells with NAPL can be used to collect water 
samples from beneath or above the NAPL. 


C3-31 


C.3.2.1.2 Collecting Mixed Ground-water/NAPL Samples 

This method involves collecting a sample of ground water and NAPL from a monitoring well, 
placing the sample in a sealed container used for volatile organics analysis being careful to ensure 
there is no headspace, allowing the sample to reach equilibrium, and submitting the water above or 
below the floating NAPL to a qualified laboratory for analysis. A disposable bailer generally works 
best for collection of this type of sample. Smith et al. (1981) has information on how to conduct 
such a test for LNAPL. Two or three samples should be collected from different monitoring wells 
containing NAPL at the site. This test should only be done when it is not possible to collect a dis¬ 
crete sample from above or below the NAPL. 

C.3.2.2 Equilibrium Partitioning Calculations 

The NAPL present at a site represents a continuing source of contamination because chlorinated 
solvents, BTEX, and other compounds will partition from the NAPL into the ground water. In such 
cases, it is generally necessary to estimate the dissolved concentration of contaminants expected in 
ground water near the LNAPL. Partitioning calculations can be performed for sites with NAPL to 
quantify contaminant loading from the NAPL into the ground water at the time the ground water or 
NAPL samples are collected. Such calculations allow a crude estimation of the impact of continuing 
sources of contamination on dissolved contaminant concentrations. The results of partitioning 
calculations may show that even if the NAPL is allowed to remain in the ground, dissolved contami¬ 
nant concentrations will remain below regulatory guidelines. This is especially true when weathered 
NAPLs with initially low contaminant concentrations are present. Partitioning calculations made by 
Wiedemeier et al. (1993) showed that NAPL present in the subsurface at a fueling facility near 
Denver, Colorado, was incapable of producing dissolved contaminant concentrations in ground water 
above regulatory standards. Such partitioning calculations should be confirmed with an LTM pro¬ 
gram. 

On the other hand, if partitioning calculations indicate that continued dissolution will produce 
contaminant concentrations exceeding regulatory guidelines, further work will be needed. The 
contaminant concentrations calculated by equilibrium methods will clearly not provide mass flux 
estimates that can be used in modeling; again, the “black box” methods will be more useful. More¬ 
over, there is no estimation of the actual mass flux across the entire body of NAPL and, therefore, 
source lifetimes and weathering rates cannot be estimated directly from partitioning data. More 
advanced calculations, such as those that will be discussed in later sections, are then required, keep¬ 
ing in mind that greater uncertainties will be introduced. 

When found in the saturated zone, residual NAPL is extremely difficult to remove. Maximum 
contaminant concentrations resulting from such partitioning will occur when the ground water and 
NAPL reach equilibrium. Assuming that equilibrium is reached gives the most conservative model¬ 
ing results. 

C.3.2.2.1 Equilibrium Partitioning of Contaminants from Mobile NAPL into Ground Water 

Because most NAPLs will be a mixture of compounds, the solubilities of those compounds will 
be lower than the solubility of the individual compound (which is what is most commonly found in 
the literature). For an organic NAPL mixture, the dissolved concentration of each compound (in 
equilibrium with the mixture) can be approximated by: 

C M ^X m C mtP eq.C.3.14 

Where: 

C sat m = solubility of compound from mixture 
X m = mole fraction of compound in the mixture 
C sat p = solubility of pure compound 


C3-32 


This equilibrium concentration may also be referred to as the effective solubility of the compound 
from the mixture. Experimental evidence (Banerjee, 1984; Broholm and Feenstra, 1995) have 
suggested that eq. C.3.14 produces reasonable approximations of effective solubilities for mixtures 
of structurally similar compounds, and that the relationship works best for binary mixtures of similar 
compounds. For other mixtures, the error is greater due to the complex solubility relationships 
created; however, the method is appropriate for many environmental studies for which there are 
many other uncertainties (Feenstra and Guiguer, 1996). 

For complex mixtures (e.g., multiple identified and unidentified solvents, or mixed fuels and 
solvents), it will be necessary to estimate the weight percent and an average molecular weight of the 
unidentified fraction of the NAPL before the calculation can be completed. In doing so, it should be 
remembered that increasing the average molecular weight for the unidentified fraction will produce 
greater estimated effective solubilities for the identified contaminants. A higher molecular weight 
for the unidentified fraction will result in a lower mole fraction for that fraction and, therefore, 
higher mole fractions (and solubilities) for the known compounds. Feenstra and Guiguer (1996) 
provide an example of these calculations for a mixture of chlorinated and nonchlorinated com¬ 
pounds. 

In the case of fuel hydrocarbon mixtures, experimental partitioning data has been collected and 
used to develop individual-compound solubility calculations, largely because fuel mixtures are 
somewhat consistent in their makeup. The fuel-water partitioning coefficient, K /tv , is defined as the 
ratio of the concentration of a compound in the fuel to the compound’s equilibrium concentration in 
water in contact with the fuel: 


- eq. C.3.15 

Where: 

K ^ = fuel-water partitioning coefficient [dimensionless] 

C f = concentration of the compound in the fuel [M/L 3 ] 

C w = concentration of the compound dissolved in ground water [M/L 3 ] 

A summary of values of for BTEX and trimethylbenzenes (TMB) in jet fuel and gasoline are 
presented by Wiedemeier et al. (1995d), along with the relationships relating to the aqueous 
solubility of a pure compound in pure water, S, which can be used to estimate K for compounds for 
which there is no experimental data. 

Using the definition of presented above, the maximum (equilibrium) total dissolved BTEX 
concentration resulting from the partitioning of BTEX from NAPL into ground water is given by: 


„ Cf 

c « = ir- eq. C.3.16 

This relationship predicts the concentration of dissolved BTEX in the ground water if the LNAPL is 
allowed to remain in contact with the ground water long enough so that equilibrium between the two 
phases is reached. Further discussion and example calculations for this method are presented by 
Wiedemeier et al. (1995d). 

C.3.2.3 Mass Flux Calculations 

In general, the rate of mass transfer from a NAPL can be given as the product of a mass transfer 
coefficient, a concentration difference, and a contact area. As Feenstra and Guiguer (1996) note, the 
driving force for mass transfer is the concentration difference across a boundaiy layer between the 
NAPL and the ground water. The concentration difference can be approximated using the effective 
solubility of a compound (eq. C.3.14) and either the measured concentration of the compound in 
ground water adjacent to the NAPL, or a calculated (theoretical) ground-water concentration. How- 


C3-33 




ever, the contact area and the mass transfer coefficient incorporate a great deal of uncertainty and are 
typically calibration parameters for modeling dissolution, as discussed previously. 

Once these parameters have been estimated, one can use them in a variety of models. In gen¬ 
eral, models for dissolution of NAPL in porous media either assume local equilibrium between 
phases, or assume that dissolution is a first-order process governed by the variables discussed above 
(Feenstra and Guiguer, 1996). Abriola and Pinder (1985a), Baehr and Corapcioglu (1987), and 
Kaluarachchi and Parker (1990) developed two-dimensional NAPL migration models that account 
for dissolution using the local equilibrium assumption (LEA). As noted by Abriola (1996), these 
studies generally were computer modeling studies for which follow-up laboratory work is ongoing 
and uncovering additional factors to consider. For single-component NAPLs, models utilizing a 
first-order reaction have been developed by Miller et al. (1990), Powers et al. (1992), Brusseau 
(1992), Guiguer (1993), and Guiguer and Frind (1994). For multi-component NAPLs, a model 
developed by Shiu et al. (1988) and Mackay et al. (1991) may be of use. 

Due to approximate nature of flux calculations and the inherent uncertainty in those calcula¬ 
tions, we have chosen to omit a detailed discussion of such efforts. The numerical modeling using 
LEA methods is beyond the scope of this work, and may not be practical for use at most sites. In¬ 
stead, we will present a brief review of ideas presented by Feenstra and Guiguer (1996) and Johnson 
and Pankow (1992) in order to illustrate some of the concepts involved in estimating flux terms. 
Should further detail or other methods be desired, both of those works provide excellent background 
and references to start with, including many of the works referenced in this discussion of source term 
calculations. 

C.3.2.3.1 General Mass Transfer Models 

Using concepts from the field of chemical engineering, Feenstra and Guiguer (1996) note that 
for a single-component NAPL, simple dissolution of the compound may be described by: 

N=K c {C w -C sal ) eq.C.3.17 

Where: 

N = flux of the species of interest (M/L 2 T) 

K c = mass transfer coefficient (L/T) 

C H = concentration of compound in bulk aqueous phase (M/L 3 ) 

C sat = concentration of compound at NAPL-water interface (taken as the solubility of the 
compound) (M/L 3 ) 

The mass transfer coefficient may be calculated various ways, but in all cases, the diffusivity of the 
species of interest is a factor. Feenstra and Guiguer (1996) present three methods for determining a 
mass transfer coefficient. 

In a porous media, the mass transfer rate per volume of porous medium can be defined by 
multiplying the mass flux by the ratio of NAPL surface contact area to the unit volume of porous 
medium, yielding: 

N' = 4C„-C S „) eq.c.3.18 

Where: 

TV* - flux of the species of interest per unit volume of porous medium (M/L 2 T) 

^ = lumped mass transfer coefficient (L/T) 

C m = concentration of compound in bulk aqueous phase (M/L 3 ) 

C sat = concentration of compound at NAPL-water interface (taken as the solubility of the 
compound) (M/L 3 ) 

The lumped mass transfer coefficient is the product of K c and the ratio of the NAPL surface contact 


C3-34 


area and the unit volume of the porous media. This can further be extended for multicomponent 
NAPLs: 

< = eq.C.3.19 

Where: 

N* m - flux of component m per unit volume of porous medium (M/L 2 T) 

A m - lumped mass transfer coefficient for component m (L/T) 

C w m = concentration of component m in bulk aqueous phase (M/L 3 ) 

C sat m - concentration of component m at NAPL-water interface (calculated using eq. C.3.14) 
(M/L 3 ) 

Further complicating all of these relationships is the fact that as dissolution continues, A m will vary 
over time as the amount of NAPL changes. This can be accounted by using the following first-order 
relation: 

N„ = S„UC,',,„-C w , m ) eq.C.3.20 

Where: 

N m = flux of component m per unit volume of porous medium (M/L 2 T) 

S w = average fraction of pore volume occupied by water 
A m = lumped mass transfer coefficient for component m (L/T) 

C w m = concentration of component m in bulk aqueous phase (M/L 3 ) 

C sat m = concentration of component m at NAPL-water interface (calculated using eq. C.3.14)) 
(M/L 3 ) 

Again, it bears repeating that on the field scale, measurement of many of the parameters used 
for these calculations is not possible, and, therefore, great uncertainty is introduced. Source terms 
calculated using these or any other methods should be presented in that light, and if used for solute 
transport modeling, should be accompanied with a sensitivity analysis. 

C.3.2.3.2 Nonequilibrium Partitioning Model of Johnson and Pankow (1992) 

The steady-state, two-dimensional dissolution of contaminants from a pool of NAPL floating on 
the water table into ground water (assumed to be a semi-infinite medium) can be described by the 
steady-state, two-dimensional, advection-dispersion equation (Hunt etal ., 1988): 

ac „ e 2 c 

v—=D ! — t x,z> 0 eq.C.3.21 

OX oz 

Where: 

C = contaminant concentration dissolved in water 
= average linear ground-water velocity 
D z = vertical dispersion coefficient 
If it is assumed that: 

• The time required for total NAPL dissolution is exceedingly long in comparison to the 
contact time between the NAPL pool and the flowing ground water 

• The NAPL pool is wide compared to the horizontal transverse mixing process 

• The NAPL pool can be approximated as a rectangle 

• The NAPL lens width does not affect the dissolution rate 

• The elevation of the NAPL lens is taken as z=0, with z measured positively upward 

• The boundary conditions are: 


C3-35 




C(x, z = ) = 0 

C(x, z = 0) = C e 0 < x > L 

C(x = 0, z) = 0 

Where: 

C = contaminant concentration dissolved in water 
C e = effective water solubility 
L = horizontal length of NAPL pool 

then the rate of dissolution of constituents from an LNAPL lens into ground water flowing beneath 
the lens can be calculated as two-dimensional, steady-state dissolution, and the surface area averaged 
mass transfer rate, M , is calculated as (Johnson and Pankow, 1992; Hunt et al., 1988): 

a 


M a = Cn e 


'4 

7rL 


eq. C.3.22 


Where: 

n e = effective porosity 

L - length of NAPL lens parallel to ground-water flow direction 
= average linear ground-water flow velocity 

C = effective water solubility (proportional to a compound’s pure phase solubility and mole 
fraction in the NAPL) 

D_ = vertical dispersion coefficient 

The vertical dispersion coefficient, D z , results from a combination of molecular diffusion and me¬ 
chanical dispersion and is defined as (Johnson and Pankow, 1992): 

D z =D e +v x a z eq. C.3.23 

Where: 

D = effective molecular diffusivity (corrected for porosity and tortuosity) 
a, = vertical dispersivity (typically 0.01 of longitudinal dispersivity) 

= average linear ground-water flow velocity 

A typical value of D e for a nonpolar organic compound is 1 x 10' 5 cm 2 /sec (Sellers and Schreiber, 
1992). 

“At very low flow velocities where molecular diffusion dominates, the average concentration 
decreases with increasing flow velocity because of decreasing contact time. At higher groundwater 
flow velocities where dispersion dominates over diffusion, average percent solubility becomes 
independent of velocity. This is because the transverse dispersion coefficient is proportional to flow 
velocity, and D z /v is constant. At typical groundwater flow velocities, an effluent concentration far 
less than the solubility limit is expected. For example, for a flow velocity of 1 m/day and a.=10‘ 4 m, 
less than 1 percent of solubility is predicted, and considerable pumping would be required to remove 
the contaminant. The analysis predicts a constant contaminant concentration dissolved in the ex¬ 
tracted water as long as the separate phase covers the boundary” (Hunt et al., 1988, pp. 1253 and 
1254). 


C3-36 




C.3.3 CONFIRMING AND QUANTIFYING BIODEGRADATION 

Chemical evidence of two types can be used to document the occurrence of biodegradation. 

The first type of evidence is graphical and is provided by the electron acceptor and metabolic 
byproduct maps discussed in Section C.2. The second line of evidence involves using a conservative 
tracer. 


C.3.3.1 Isopleth Maps 

The extent and distribution of contamination relative to electron acceptors and metabolic 
byproducts can be used to qualitatively document the occurrence of biodegradation. Depleted 
dissolved oxygen concentrations in areas with fuel hydrocarbon contamination indicates that an 
active zone of aerobic hydrocarbon biodegradation is present. Depleted nitrate and sulfate concen¬ 
trations in areas with fuel hydrocarbon contamination indicate that an active zone of anaerobic 
hydrocarbon biodegradation is present and that denitrification and sulfate reduction are occurring. 
Elevated iron (II) and methane concentrations in areas with fuel hydrocarbon contamination indicate 
that an active zone of anaerobic hydrocarbon biodegradation is present and that iron reduction and 
methanogenesis are occurring. Isopleth maps of contaminants, electron acceptors, and metabolic 
byproducts can be used as evidence that biodegradation of fuel hydrocarbons is occurring. 

Figures C.2.7 and C.2.8 show how these maps can be used to support the occurrence of biodegrada¬ 
tion. Figure C.2.7 shows that areas with depleted dissolved oxygen, nitrate, and sulfate correspond 
with areas having elevated BTEX concentrations. Figure C.2.8 shows that areas with elevated 
iron (II) and elevated methane concentrations also coincide with areas having elevated BTEX con¬ 
centrations. These figures suggest that aerobic respiration, denitrification, iron reduction, sulfate 
reduction, and methanogenesis are all occurring at the example site. 


C.3.3.2 Data Set Normalization 

In order to calculate biodegradation rates accurately, measured contaminant concentrations must 
be normalized for the effects of dispersion, dilution, and sorption. A convenient way to do this is to 
use compounds or elements associated with the contaminant plume that are relatively unaffected or 
predictably affected by biologic processes occurring within the aquifer. At sites where commingled 
fuel hydrocarbon and chlorinated solvent plumes are present, the trimethylbenzene isomers (TMB), 
which can be biologically recalcitrant under some geochemical conditions have proven useful when 
estimating biodegradation rates for BTEX and chlorinated solvents. At sites where TMB data are 
not available, the chloride produced as a result of biodegradation or the carbon nucleus of the chlori¬ 
nated compound can be used as a tracer. 

Measured concentrations of tracer and contaminant from a minimum of two points along a flow 
path can be used to estimate the amount of contaminant that would be expected to remain at each 
point if biodegradation were the only attenuation process operating to reduce contaminant concentra¬ 
tions. The fraction of contaminant remaining as a result of all attenuation processes can be com¬ 
puted from the measured contaminant concentrations at two adjacent points. The fraction of con¬ 
taminant that would be expected to remain if dilution and dispersion were the only mechanisms for 
attenuation can be estimated from the tracer concentrations at the same two points. The tracer is 
affected by dilution and dispersion to the same degree as the contaminant of interest and is not 
affected by biologic processes. The following equation uses these assumptions to solve for the 
expected downgradient contaminant concentration if biodegradation had been the only attenuation 
process operating between two points along the flow path: 


CB,corr ~ ^B 


f T \ 

L A 

\T B J 


eq. C.3.24 


C3-37 




Where: 

C B corr = corrected contaminant concentration at a point B downgradient 
C B = measured contaminant concentration at point B 
T a = tracer concentration at a point A upgradient 
T b = tracer concentration at point B downgradient 

This equation can be used to estimate the theoretical contaminant concentration that would result 
from biodegradation alone for every point along a flow path on the basis of the measured contami¬ 
nant concentration at the origin and the dilution of the tracer along the flow path. This series of 

normalized concentrations can then be used to estimate a first-order rate of biodegradation as de¬ 
scribed in Section C.3.3.3. 

C.3.3.2.1 Normalization Using Organic Compounds as Tracers 

A convenient way of estimating biodegradation rate constants is to use compounds present in 
the dissolved contaminant plume that that are biologically recalcitrant. One such compound that is 
useful in some, but not all, ground-water environments is Trimethylbenzene (TMB). The three 
isomers of this compound (1,2,3-TMB, 1,2,4- TMB, and 1,3,5- TMB) are generally present in suffi¬ 
cient quantities in fuel mixtures to be readily detectable when dissolved in ground water. When 
chlorinated solvents enter the subsurface as a mixture with petroleum hydrocarbons, the TMB 
compounds can be useful tracers. The TMB isomers are fairly recalcitrant to biodegradation under 
anaerobic conditions; however, the TMB isomers do not make good tracers under aerobic conditions 
(because they are readily biodegraded in aerobic environments). The degree of recalcitrance of TMB 
is site-specific, and the use of this compound as a tracer must be evaluated on a case-by-case basis. 
Nevertheless, if any TMB mass is lost to biodegradation, equation C.3.24 will be conservative 
because the calculated mass losses and the attenuation rate constants calculated on the basis of those 
losses will be lower than the actual losses and attenuation rates. Another compound of potential use 
as a conservative tracer is tetramethylbenzene; however, detectable dissolved tetramethylbenzene 
concentrations are generally less common than detectable dissolved TMB concentrations. 

An ideal tracer would have Henry’s Law and soil sorption coefficients identical to the contami¬ 
nant of interest; however, TMB is more hydrophobic than BTEX, chlorinated ethenes, and chlori¬ 
nated ethanes, resulting in a higher soil sorption coefficient than the compound of interest. As a 
result, use of TMB as a tracer is often conservative, and the biodegradation rates can be underesti¬ 
mated. It is best, whenever possible, to compare several tracers to determine whether they are 
internally consistent. 

C.3.3.2.2 Normalization Using Inorganics as Tracers 

Inorganic compounds also can serve as tracers for the contaminant of interest as long as their 
presence is in some way associated (either directly or indirectly) with the dissolved contaminant 
plume. For many chlorinated solvent plumes, the sum of ionic chloride and organic chloride associ¬ 
ated with the solvents can be considered a conservative tracer. Note that the following discussion 
assumes that the background chloride concentration is negligible in comparison to the source area 
concentration of total chloride plus chlorine. If background chloride is more than approximately 10 
percent of the total source area chloride plus chlorine concentration, then background concentrations 
will need to be accounted for prior to performing the tracer normalization. 

Total chlorine can easily be calculated by multiplying the measured concentration of a chlori¬ 
nated organic compound by the mass fraction of chlorine in the molecule, then summing that quan¬ 
tity for all the chlorinated organic compounds represented in the plume. The stoichiometry for 
chlorinated ethenes is presented in the following paragraphs. 


C3-38 


As PCE is reduced to ethene, 4 moles of chloride are produced: 

C 2 C1 4 ->C 2 H 4 + 4C1- 

On a mass basis, the ratio of chloride produced to PCE degraded is given by: 

Molecular weights: PCE 2(12.011) + 4(35.453) = 165.83 gm 

Chloride 4(35.453) = 141.81 gm 

Mass Ratio of Chloride to PCE = 141.81:165.83 = 0.86:1 

Similarly, as TCE is reduced to ethene, 3 moles of chloride are produced: 

C 2 C1 3 H->C 2 H 4 + 3C1 

On a mass basis, the ratio of chloride produced to TCE degraded is given by: 

Molecular weights: TCE 2(12.011) + 3(35.453) + 1(1.01)= 131.39 gm 

Chloride 3(35.453) = 106.36 gm 

Mass Ratio of Chloride to TCE = 106.36:131.39 = 0.81:1 

Likewise, as DCE is reduced to ethene, 2 moles of chloride are produced: 

C 2 C1 2 H 2 ->C 2 H 4 + 20' 

On a mass basis, the ratio of chloride produced to DCE degraded is given by: 

Molecular weights: DCE 2(12.011) + 2(35.453) + 2(1.01)= 96.95 gm 

Chloride 2(35.453) = 70.9 gm 

Mass Ratio of Chloride to DCE = 70.9:96.95 = 0.73:1 

As VC is reduced to ethene, 1 mole of chloride is produced: 

C,C1H,—»C,H, + Cl 

2 3 2 4 

On a mass basis, the ratio of chloride produced to VC degraded is given by: 

Molecular weights: VC 2(12.011) + 1(35.453) + 3(1.01)= 62.51 gm 

Chloride 1(35.453) = 35.453 gm 

Mass Ratio of Chloride to VC = 35.453:62.51 = 0.57:1 

Therefore, the amount of total chloride plus chlorine for a spill undergoing reductive dechlorination 
would be estimated as: 

[Cl Tot J = 0.86[PCE] + 0.81 [TCE] + 0.73[DCE]) + 0.57[VC]) eq. C.3.25 

Example C.3.4: Calculating Total Concentration of Chloride and Organic Chlorine 

The approach is illustrated in the following data set from the West TCE Plume at the St. Joseph, 
Michigan NPL site. 

A series of discrete vertical water samples were taken in transects that extended across the 
plume at locations downgradient of the source of TCE. The locations of the samples are depicted in 
Figure C.3.5 as circles. At each sampling location, water samples were acquired using a hollow- 
stem auger. The leading auger was slotted over a five-foot interval. After a sample was collected, 
the auger was driven five feet further into the aquifer and the next sample was collected. At any one 
location, the water samples were collected in a sequential and contiunuous series that extended from 
the water table to a clay layer at the bottom of the aquifer. The concentrations of contaminants at 
each location were averaged in water samples that extend across the entire vertical extent of the 
plume. The location with the highest average concentration of chlorinated ethenes in a particular 
transect was selected to represent the centerline of the plume. The locations of the sample locations 
in the centerline of the plume are depicted in Figure C.3.5 as open circles. Each centerline location 
is labelled in an oval. 


C3-39 




Lake Michigan 


Transect 5 


Transect 4 


Transect 1 

• A 


588 


Transect 2 


592 




594 


596 


598 


Source 


FEET 


Figure C.3.4 Location of sampling points at the St. Joseph, Michigan, NPL site. 


The concentrations of chlorinated ethenes and chloride in the centerline of the TCE plume at St. 
Joseph, Michigan, are presented in Table C.3.5. 

Table C.3.5 Attenuation of Chlorinated Ethenes and Chloride Downgradient of the Source of TCE in the 

West Plume at the St. Joseph, Michigan, NPL Site. 


Compound 

Sampling Locations 


T-2-5 T-l-4 

T-4-2 

T-5-3 

55AE 


Distance Downgradient (feet) 


0 

200 

1.000 

1.500 

2.000 


Tmcr/T itpr't 


\ lllw 1—/1ICI / ----- - - - ---- 

PCE 

0.0 

0.0 

0.0 

0.0 

0.0 

TCE 

12.1 

3.4 

1.3 

0.035 

0.022 

Total DCE 

37.6 

11.7 

2.4 

0.23 

0.45 

Vinyl 

Chloride 

2.3 

3.7 

0.51 

0.063 

0.070 

Total 

Organic 

Chloride 

38.5 

13.4 

3.2 

0.2 

0.4 

Chloride 

89.7 

78.6 

98.9 

63.6 

54.7 

Tracer (Total 
Chloride plus 
Chlorine) 

128.2 

92.0 

102.1 

63.8 

55.1 


C3-40 




































At the monitoring point closest to the source of the plume (see location T-2-5 in Table C.3.5 and 
Figure C.3.4) the concentrations of TCE, total DCE, vinyl chloride and chloride were 12.1, 37.6, 2.3 
and 89.7 mg/L, respectively. This results in an upgradient tracer concentration of 


TCE chlorine 

+ 

(0.809)(12.1 mg/L) 

DCE chlorine 

+ 

(0.731)(37.6 mg/L) 

Vinyl chloride chlorine 

+ 

(0.567)(2.3 mg/L) 

Chloride 

+ 

89.7 mg/L 

Total chloride plus chlorine 

= 

128.2 mg/L 


At the downgradient location 55AE, which is 2,000 feet from the source, the concentrations of 
TCE, total DCE, vinyl chloride, and chloride were 0.022, 0.45, 0.070, and 54.70 mg/L, respectively. 
This results in a downgradient concentration of 


TCE chlorine 

+ 

(0.809)(0.022 mg/L) 

DCE chlorine 

+ 

(0.731)(0.45 mg/L) 

Vinyl chloride chlorine 

+ 

(0.567)(0.070 mg/L) 

Chloride 

+ 

54.7 mg/L 

Total chloride plus chlorine 

— 

55.1 mg/L 


The computed series of total chloride plus chlorine concentrations can be used with equation 
C.3.24 to estimate a normalized data set for contaminant concentrations. 

Example C.3.5 : Normalizing Contaminant Concentrations Along a Flow Path 

Equation 3.24 will be used to calculate a normalized concentration for TCE at the locations 
depicted in Figure C.3.4 and Table C.3.5. Given are the observed concentrations of TCE and tracer 
(Table C.3.5) for five points that form a line parallel to the direction of ground-water flow (Figure 
C.3.4) To calculate normalized concentrations of TCE using the attenuation of the tracer, the dilu¬ 
tion of the tracer is caculated at each location by dividing the concentration of tracer at the source (or 
most contaminated location) by the concentration of tracer at each downgradient location. Then the 
measured concentration of TCE downgradient is multiplied by the dilution of the tracer. The cor¬ 
rected concentrations of TCE are presented in Table C.3.6. This information will be used in sections 
C.3.3.3 to calculate the rate of natural biodegradation of TCE. 

Table C.3.6* Use of the Attenuation of a Tracer to Correct the Concentration of TCE Downgradient of the 
Source of TCE in the West Plume at the St. Joseph, Michigan, NPL Site 


Compound 

Sampling Locations 


T-2-5 

T-l-4 | T-4-2 1 T-5-3 

55AE 


Distance Down Gradient (feet) 


0 200 1.000 1.500 

2.000 


itor\ _____ 

..... ............. ^ / -- 

TCE 

12.1 

3.4 

1.3 

0.035 

0.022 

Tracer 

128.2 

92.0 

102.1 

63.8 

55.1 

Dilution of Tracer 

128.2/28.2 

128.2/92.0 

128.2/ 102.1 

128.2/ 63.8 

128.2/55.1 

Corrected TCE 

12.1 

4.7 

1.6 

0.070 

0.051 


C.3.3.3 Calculating Biodegradation Rates 

Several methods, including first- and second-order approximations, may be used to estimate the 
rate of biodegradation of chlorinated compounds when they are being used to oxidize other organic 
compounds. Use of the first-order approximation can be appropriate to estimate biodegradation rates 


C3-41 

























for chlorinated compounds when the rate of biodegradation is controlled solely by the concentration 
of the contaminant. However, the use of a first-order approximation may not be appropriate when 
more than one substrate is limiting microbial degradation rates or when microbial mass is increasing 
or decreasing. In such cases, a second- or higher-order approximation may provide a better estimate 
of biodegradation rates. 

C.3.3.3.1 First-Order Decay 

As with a large number of processes, the change in a solute’s concentration in ground water 
over time often can be described using a first-order rate constant. A first-order approximation, if 
appropriate, has the advantage of being easy to calculate and simplifies fate and transport modeling 
of complex phenomenon. In one dimension, first-order decay is described by the following ordinary 
differential equation: 



eq. C.3.26 


Where: 


C = concentration at time t [M/L 3 ] 
k = overall attenuation rate (first-order rate constant) [1/T] 


Solving this differential equation yields: 



eq. C.3.27 


The overall attenuation rate groups all processes acting to reduce contaminant concentrations and 
includes advection, dispersion, dilution from recharge, sorption, and biodegradation. To determine 


the portion of the overall attenuation that can be attributed to biodegradation, these effects must be 


accounted for, and subtracted from the total attenuation rate. 

Aronson and Howard (1997) have compiled a large number of attenuation rate constants for 


biodegradation of organic compounds in aquifers. This information is supplied to provide a basis for 
comparison of rate constants determined for at a particular site to the general experience with natural 
attenuation as documented in the literature. It is not intended to provide rate constants for a site in a 
risk assessment or exposure assessment. The rate constants used to describe behavior of a particular 


site must be extracted from site characterization information particular to that site. 

The distribution of the rate constants reported for TCE is presented in Figure C.3.5. Notice that 
the average rate is near 1.0 per year, and that most of the rates cluster in a relatively narrow range 
between 3.0 per year and 0.3 per year. Some of the published rates are very low, less than 0.1 per 
year. The report compiles data from sites where rates are published. The general bias against pub¬ 
lishing negative data suggests that there are many plumes where TCE attenuation was not detectable 
(Type 3 behavior), and that data on these plumes is not found in the literature. The data from 
Aronson and Howard (1997) reflect the behavior of plumes where reductive dechlorination is an 
important mechanism (Type 1 and Type 2 sites). Rate constants for PCE and Vinyl Chloride are 
presented in Figures C.3.6 and C.3.7. The average rate for dechlorination of PCE is somewhat faster 
than for TCE, near 4.0 per year, and the rate for Vinyl Chloride is slower, near 0.6 per year. 

Two methods for determining first-order biodegradation rates at the field scale are presented. 
The first method involves the use of a normalized data set to compute a decay rate. The second 
method was derived by Buscheck and Alcantar (1995) and is valid for steady-state plumes. 
Wiedemeier et al. (1996b) compare the use of these two methods with respect to BTEX biodegrada¬ 
tion. 


C3-42 




Figure C.3.5. Field rate constants for TCE as reported in literature. 



Figure C.3.6 Field rate constants for PCE as reported in literature. 


C3-43 


Rate Constant (/yr) Ra,e Constant ^ 









































100 



Figure C.3.7 Field rate constants for vinyl chloride as reported in literature. 

C.3.3.3.2 Use of a Normalized Data Set 

In order to ensure that observed decreases in contaminant concentrations can be attributed to 
biodegradation, measured contaminant concentrations must be corrected for the effects of advection, 
dispersion, dilution from recharge, and sorption, as described in Section C.3.3.2 using equation 
C.3.24. The corrected concentration of a compound is the concentration that would be expected at 
one point (B) located downgradient from another point (A) if the processes of dispersion and dilution 
had not been occurring between points A and B. 

The biodegradation rate can be estimated between any two points (A and B) of a normalized 
data set (where point A is upgradient of point B) by substituting the concentration at point A for C 0 , 
and the normalized concentration at point B, C Bcorr , for C in equation C.3.27. The resulting relation¬ 
ship is expressed as: 

c „, corr = c A e-^ eq.C.3.28 

Where: 

C Bc0 rr ” normalized contaminant concentration at downgradient point B (from eq. C.3.25) 

C 4 = contaminant concentration at upgradient point A that if point A is the first point in 
the normalized data set, then C = C, 

A = first-order biological decay rate (first-order rate constant) [1/T] 
t = time of contaminant travel between points A and B 

The rate constant in this equation is no longer the total attenuation rate, k, but is the biological 
decay rate, A, because the effects of advection, dispersion, dilution from recharge, and sorption have 
been removed (Section C.3.3.2). This relationship can be used to calculate the first-order biological 
decay rate constant between two points by solving equation C.3.28 for A: 


C3-44 











f 


In 


C 


\ 


B,corr 

V C A J 


The travel time, t, between two points is given by: 


t X 

V c 

Where: 

x = distance between two points [L] 
v c = retarded solute velocity [L/T] 


eq. C.3.29 


eq. C.3.30 


Example C.3.6 : First-Order Decay Rate Constant Calculation Using Normalized Data Set 

Equation C.3.30 and C.3.29 can be used to calculate rate constants between any two points 
along a flow line. For travel from locations T-2-5 and and 55AE in Figure C.3.4 and Table C.3.6, 
the upgradient concentration of TCE is 12.1 mg/1, the corrected downgradient concentration is 0.051 
mg/1, and the distance between the locations is 2,000 feet. 

From Figure C.3.4, the water table drops 10 feet as the plume moves 1,300 feet from transect 1 
to transect 5. The site has a hydraulic gradient of 0.008 feet per foot. Aquifer testing at the site 
predicts an average hydraulic conductivity of 50 feet per day. If the effecive porosity of the sandy 
aquifer is assumed to be 0.3, the seepage velocity (V ) would be (Equation C.3.6): 

= 0.4 ft / day x 0.008 ft/ft = 

0.3 


The average organic matter content of the aquifer matrix material is less than the detection limit 
of 0.001 g/g. We will assume the organic matter content is equal to the detection limit. If the K oc of 
TCE is 120 ml/g, the porosity is 0.3, and the bulk density is 1.7 gm/cm 3 , the distribution of TCE 
between ground water and aquifer solids is the product of the K oc , the fraction organic carbon, the 
bulk density, divided by the porosity, or 0.3. The retarded velocity of TCE compared to water (R) 
would be (Equation C.3.8 and Equation C.3.13): 

R = 1 + 120 (ml/g) * 0.001 (g/g) * 1.7 (g/cm 3 )/ 0.3 (ml/ ml) = 1.7 

The velocity of TCE in the aquifer would be equal to the velocity of water in the plume divided by 
the retardation of TCE. The TCE velocity (v c ) would be: 

v c = 1.3 feet per day/1.7 = 0.8 feet per day 

If the distance between the wells is 2,000 feet, and the retarded velocity of TCE is 0.8 feet per 
day, by equation C.3.30 the travel time is: 

t = 2,000 feet/ 0.8 feet per day = 2,500 days = 6.8 years 
From equation C.3.29, the rate of biotransformation between locations T-2-5 and 55 AE is: 

X = In (0.055/12.1)/ 6.8 per year = 0.79 per year 

If a number of sampling locations are available along a flow path, all the locations should be 
included in the calculation of the biotransformation rate. The simplest way to determine the first- 
order rate constant from an entire set of normalized data is to make a log-linear plot of normalized 
contaminant concentrations versus travel time. If the data plot along a straight line, the relationship 
is first-order and an exponential regression analysis can be performed. 

The exponential regression analysis gives the equation of the line of best fit for the data being 
regressed from a log-linear plot and has the general form: 


C3-45 








Where: 


y = be mx 


eq. C.3.31 


y = y axis value 
b - y intercept 
m = slope of regression line 
x - x-axis value 

When using normalized data, x is the contaminant travel time to the downgradient locations and m is 
the first-order rate of change equal to the negative. The correlation coefficient, R 2 , is a measure of 
how well the regression relationship approximates the data. Values of R 2 can range from 0 to 1; the 
closer R 2 is tol, the more accurate the equation describing the trend in the data. Values of R 2 greater 
than 0.80 are generally considered useful; R 2 values greater than 0.90 are considered excellent. 
Several commonly available spreadsheets can be used to facilitate the exponential regression analy¬ 
sis. The following example illustrates the use of this technique. 

Figure C.3.8 depicts a regression of normalized TCE concentration against travel time 
downgradient. The slope of the exponential regression is -0.824* where * is travel time in years, 
corresponding to a first-order rate of change of-0.824 per year and a first-order rate of biodegrada¬ 
tion of 0.824 per year. In Figure C.3.8, an exponential regression was performed on the normalized 
concentrations of TCE against time of travel along the flow path. An alternative approach would be 
to perform a linear regression of the natural logarithm of the normalized concentration of TCE 
against travel time along the flow path. 


Travel Distance (feet) 


500 1000 1500 2000 



Figure C.3.8 Exponential regression of TCE concentration on time of travel along flow path. 


C3-46 










C.3.3.3.3. Method of Buscheck and Alcantar (1995) 

Buscheck and Alcantar (1995) derive a relationship that allows calculation of first-order decay 
rate constants for steady-state plumes. This method involves coupling the regression of contaminant 
concentration (plotted on a logarithmic scale) versus distance downgradient (plotted on a linear 
scale) to an analytical solution for one-dimensional, steady-state, contaminant transport that includes 
advection, dispersion, sorption, and biodegradation. For a steady-state plume, the first-order decay 
rate is given by (Buscheck and Alcantar, 1995): 


fr 


A-- 


4 


V L 


1 + 2 ar 


( l V 


\ V x J 


2 \ 

-1 

J 


eq. C.3.32 


Where: 

A = first-order biological rate constant 
v = retarded contaminant velocity in the x-direction 
a x = dispersivity 

k/v x - slope of line formed by making a ln-linear plot of contaminant concentration versus 
distance downgradient along flow path 


Exa mple C.3.7 : First-Order Rate Constant Calculation Using Method of Buscheck and Alcantar 
(1995) 

The first step is to confirm that the contaminant plume has reached a steady-state configuration. 
This is done by analyzing historical data to make sure that the plume is no longer migrating 
downgradient and that contaminant concentrations are not changing significantly through time. This 
is generally the case for older spills where the source has not been removed. The next step is to 
make a plot of the natural logarithm of contaminant concentration versus distance downgradient (see 
Figure C.3.9). Using linear regression, y in the regression analysis is the contaminant concentration, 
x is the distance downgradient from the source, and the slope of the ln-linear regression is the ratio k/ 
v v that is entered into equation C.3.32. 

The slope is -0.0028 feet. As calculated above, the retarded TCE velocity in the plume v c is 0.8 
feet per day. If a x = 5% of the plume length, then a x = 100 feet. Inserting these values for a x , A/v , 
and v c into equation C.3.32, the estimated value of A = -0.0016 per day or-0.59 per year. 

C. 3.3.2.2.3 Comparison of First-Order Methods 

If the data are available, concentrations of tracers should be used to normalize concentrations of 
contaminants prior to calculation of rate constants. If tracer data is not available, the method of 
Buscheck and Alcantar (1995) can be used if a value for longitudinal dispersion is available, or if 
one is willing to assume a value for longitudinal dispersion. Whenever possible, more than one 
tracer should be used to normalize the concentrations of contaminants. If the normalized concentra¬ 
tions agree using several different tracers, the approach can be accepted with confidence. In addition 
to chloride and trimethylbenzene, methane, and total organic carbon dissolved in ground water are 
often useful tracers in plumes of chlorinated solvents undergoing natural attenuation. 


C3-47 











100 


Figure C.3.9 



0 500 1000 1500 2000 2500 

Distance Downgradient (feet) 


Regression of the TCE concentration on distance along flow path. 


C3-48 









C.3.4 DESIGN, IMPLEMENTATION, AND INTERPRETATION OF MICROCOSM 
STUDIES 

C.3.4.1 Overview 

If properly designed, implemented, and interpreted, microcosm studies can provide very con¬ 
vincing documentation of the occurrence of intrinsic bioremediation. They are the only “line of 
evidence” that allows an unequivocal mass balance on the biodegradation of environmental contami¬ 
nants. If the microcosm study is properly designed, it will be easy for decision makers with non¬ 
technical backgrounds to interpret. The results of a microcosm study are strongly influenced by the 
nature of the geological material submitted to study, by the physical properties of the microcosm, by 
the sampling strategy, and the duration of the study. In addition, microcosm studies are time con¬ 
suming and expensive. A microcosm study should only be undertaken at sites where there is consid¬ 
erable uncertainty concerning the biodegradation of contaminants based on soil and ground-water 
samples alone. 

Material for a microcosm study should not be selected until the geochemical behavior of the site 
is well understood. Contaminant plumes may consume oxygen, nitrate, or sulfate, and produce iron 
(II), manganese (II), or methane. These processes usually operate concurrently in different parts of 
the plume. Regions where each process prevails may be separated in directions parallel to ground- 
water flow by hundreds of meters, in directions perpendicular to ground-water flow by tens of 
meters, and vertically by only a few meters. Rate constants and constraints for petroleum hydrocar¬ 
bon biodegradation will be influenced by the prevailing geochemistry. Material from microcosms 
must be acquired for depth intervals and locations that have been predetermined to be representative 
of the prevailing geochemical milieu in the plume. 

Contaminant biodegradation supported by oxygen and nitrate cannot be adequately represented 
in microcosm. In the field, organisms that use oxygen or nitrate proliferate until they become limited 
by the supply of electron acceptor. After that time, the rate of hydrocarbon degradation is controlled 
by the supply of electron acceptor through diffusion or hydrodynamic dispersion. Microcosms have 
been used successfully to simulate sulfate-reducing, iron-reducing, and methanogenic regions of 
plumes. Oxygen is toxic to sulfate-reducing and methanogenic microorganisms. Material should be 
collected and secured in a manner that precludes oxygenation of the sample. 

Batch microcosms that are sacrificed for each analysis usually give more interpretable results 
than column microcosms or batch microcosms that are sampled repetitively. For statistical reasons, 
at least three microcosms should be sampled at each time interval. If one assumes a first-order rate 
law, and no lag, a geometrical time interval for sampling should be the most efficient. An example 
would be sampling after 0 weeks, 2 weeks, 1 month, 2 months, 4 months, and 8 months. As a 
practical matter, long lags frequently occur, and the rate of bioremediation after the lag is rapid. A 
simple linear time scale is most likely to give interpretable results. 

The batch microcosms should have approximately the same ratio of solids to water as the 
original material. Most of the microbes are attached to solids. If a microcosm has an excess of 
water, and the contaminant is mostly in the aqueous phase, the microbes must process a great deal 
more contaminant to produce the same relative change in the contaminant concentration as would be 
obtained at field scale. The kinetics at field scale would be underestimated. 

Microcosms are inherently time consuming. At field scale, the residence time of a plume may 
be several years to decades. Slow rates of transformation may have a considerable environmental 
significance. A microcosm study that lasts only a few weeks or months may not have the resolution 
to detect slow changes that are still of environmental significance. Further, microcosms often show a 
pattern of sequential utilization, with toluene and the xylenes degrading first, and benzene and 
ethylbenzene degrading at a later time. Degradation of benzene or ethylbenzene may be delayed by 
as much as a year. 


C3-49 


As a practical matter, batch microcosms with an optimal solids-to-water ratio, sampled every 2 
months in triplicate for up to 18 months, can resolve biodegradation from abiotic losses with a rate 
detection limit of 0.001 to 0.0005 per day. Many plumes show significant attenuation of contamina¬ 
tion at field-calibrated rates that are slower than the detection limit of today’s microcosm technology. 
The most appropriate use of microcosms is to document that contaminant attenuation is largely a 
biological process. Rate constants for modeling purposes are more appropriately acquired from 
field-scale studies. 

Microcosm studies are often used to provide a third line of evidence. The potential for biodeg¬ 
radation of the contaminants of interest can be confirmed by the use of microcosms, through com¬ 
parison of removals in the living treatments with removals in the controls. Microcosm studies also 
permit an absolute mass balance determination based on biodegradation of the contaminants of 
interest. Further, the appearance of daughter products in the microcosms can be used to confirm 
biodegradation of the parent compound. 

C.3.4.2 When to Use Microcosms 

There are two fundamentally different applications of microcosms. They are frequently used in 
a qualitative way to illustrate the important processes that control the fate of organic contaminants. 
They are also used to estimate rate constants for biotransformation of contaminants that can be used 
in a site-specific transport and fate model of a plume of contaminated groundwater. This paper only 
discusses microcosms for the second application. 

Microcosms should be used when there is no other way to obtain a rate constant for attenuation 
of contaminants, in particular, when it is impossible to estimate the rate of attenuation from data 
from monitoring wells in the plume of concern. There are situations where it is impossible to com¬ 
pare concentrations in monitoring wells along a flow path due to legal or physical impediments. In 
many landscapes, the direction of ground-water flow (and water table elevations in monitoring wells) 
can vary over short periods of time due to tidal influences or changes in barometric pressure. The 
direction of ground-water flow may also be affected by changes in the stage of a nearby river or 
pumping wells in the vicinity. These changes in ground-water flow direction do not allow simple 
snap-shot comparisons of concentrations in monitoring wells because of uncertainties in identifying 
the flow path. Rate constants from microcosms can be used with average flow conditions to estimate 
attenuation at some point of discharge or point of compliance. 

C.3.4.3 Application of Microcosms 

The primary objective of microcosm studies is to obtain rate constants applicable to average 
flow conditions. These average conditions can be determined by continuous monitoring of water 
table elevations in the aquifer being evaluated. The product of the microcosm study and the continu¬ 
ous monitoring of water table elevations will be a yearly or seasonal estimate of the extent of attenu¬ 
ation along average flow paths. Removals seen at field scale can be attributed to biological activity. 
If removals in the microcosms duplicate removal at field scale, the rate constant can be used for risk 
assessment purposes (B.H. Wilson et al , 1996; Bradley, et al., 1998). 

C.3.4.4 Selecting Material for Study 

Prior to choosing material for microcosm studies, the location of major conduits of ground- 
water flow should be identified and the geochemical regions along the flow path should be deter¬ 
mined. The important geochemical regions for natural attenuation of chlorinated aliphatic hydrocar¬ 
bons are regions that are actively methanogenic; regions that exhibit sulfate reduction and iron 
reduction concomitantly; and regions that exhibit iron reduction alone. The pattern of biodegrada¬ 
tion of chlorinated solvents varies in different regions. Vinyl chloride tends to accumulate during 
reductive dechlorination of TCE or PCE in methanogenic regions (Weaver et al ., 1995; J.T. Wilson 
et al ., 1995); it does not accumulate to the same extent in regions exhibiting iron reduction and 


C3-50 


sulfate reduction (Chapelle, 1996). In regions showing iron reduction alone, vinyl chloride is con¬ 
sumed but dechlorination of PCE, TCE, or DCE may not occur (Bradley and Chapelle, 1996; 1997). 
Core material from each geochemical region in major flow paths represented by the plume must be 
acquired, and the hydraulic conductivity of each depth at which core material is acquired must be 
measured. If possible, the microcosms should be constructed with the most transmissive material in 
the flow path. 

Several characteristics of ground water from the same interval used to collect the core material 
should be determined. These characteristics include temperature, redox potential, pH, and concen¬ 
trations of oxygen, sulfate, sulfide, nitrate, iron II, chloride, methane, ethane, ethene, total organic 
carbon, and alkalinity. The concentrations of compounds of regulatory concern and any breakdown 
products for each site must be determined. The ground water should be analyzed for methane to 
determine if methanogenic conditions exist and for ethane and ethene as daughter products from 
reductive dechlorination of PCE and TCE. A comparison of the ground-water chemistry from the 
interval where the cores were acquired to that in neighboring monitoring wells will demonstrate if 
the collected cores are representative of that section of the contaminant plume. 

Reductive dechlorination of chlorinated solvents requires an electron donor to allow the process 
to proceed. The electron donor could be soil organic matter, low molecular weight organic com¬ 
pounds (lactate, acetate, methanol, glucose, etc.), H 2 , or a co-contaminant such as landfill leachate or 
petroleum compounds (Bouwer, 1994; Sewell and Gibson, 1991; Klecka et al ., 1996). In many 
instances, the actual electron donor(s) may not be identified. 

Several characteristics of the core material should also be evaluated. The initial concentration 
of the contaminated material (on a mass per mass basis) should be identified prior to construction of 
the microcosms. Also, it is necessary to know if the contamination is present as a nonaqueous phase 
liquid (NAPL) or in solution. A total petroleum hydrocarbon (TPH) analysis will determine if any 
hydrocarbon-based oily materials are present. The water-filled porosity is a parameter generally used 
to extrapolate rates to the field. It can be calculated by comparing wet and dry weights of the aquifer 
material. 

To insure sample integrity and stability during acquisition, it is important to quickly transfer the 
aquifer material into ajar, exclude air by adding ground water, and seal the jar without headspace. 
The material should be cooled during transportation to the laboratory. Incubate the core material at 
the ambient ground-water temperature in the dark before the construction of microcosms. 

At least one microcosm study per geochemical region should be completed. If the plume is 
over one kilometer in length, several microcosm studies per geochemical region may need to be 
constructed. 

C.3.4.5 Geochemical Characterization of the Site 

The geochemistry of the subsurface affects behavior of organic and inorganic contaminants, 
inorganic minerals, and microbial populations. Major geochemical parameters that characterize the 
subsurface encompasses (1) pH; (2) ORP; (3) alkalinity; (4) physical and chemical characterization 
of the solids; (5) temperature; (6) dissolved constituents, including electron acceptors; and (7) micro¬ 
bial processes. The most important of these in relation to biological processes are redox potential, 
alkalinity, concentration of electron acceptor, and chemical nature of the solids. 

Alkalinity: Field indications of biologically active portions of a plume may be identified by 
increased alkalinity, compared to background wells, from carbon dioxide due to biodegradation of 
the pollutants. Increases in both alkalinity and decrease in pH have been measured in portions of an 
aquifer contaminated by gasoline undergoing active utilization of the gasoline components 
(Cozzarelli et al ., 1995). Alkalinity can be one of the parameters used when identifying where to 
collect biologically active core material. 


C3-51 



pH: Bacteria generally prefer a neutral or slightly alkaline pH level with an optimum pH range 
for most microorganisms between 6.0 and 8.0; however, many microorganisms can tolerate a pH 
range of 5.0 to 9.0. Most ground waters in uncontaminated aquifers are within these ranges. Natural 
pH values may be as low as 4.0 or 5.0 in aquifers with active oxidation of sulfides, and pH values as 
high as 9.0 may be found in carbonate-buffered systems (Chapelle, 1993). However, pH values as 
low as 3.0 have been measured for ground waters contaminated with municipal waste leachates 
which often contain elevated concentrations of organic acids (Baedecker and Back, 1979). In ground 
waters contaminated with sludges from cement manufacturing, pH values as high as 11.0 have been 
measured (Chapelle, 1993). 

ORP: The ORP of ground water is a measure of electron activity that indicates the relative 
ability of a solution to accept or transfer electrons. Most redox reactions in the subsurface are 
microbially catalyzed during metabolism of native organic matter or contaminants. The only ele¬ 
ments that are predominant participants in aquatic redox processes are carbon, nitrogen, oxygen, 
sulfur, iron, and manganese (Stumm and Morgan, 1981). The principal oxidizing agents in ground 
water are oxygen, nitrate, sulfate, manganese (IV), and iron (III). Biological reactions in the subsur¬ 
face both influence and are affected by the redox potential and the available electron acceptors. The 
redox potential changes with the predominant electron acceptor, with reducing conditions increasing 
through the sequence oxygen, nitrate, iron, sulfate, and carbonate. The redox potential decreases in 
each sequence, with methanogenic (carbonate as the electron acceptor) conditions being most reduc¬ 
ing. The interpretation of redox potentials in ground waters is difficult (Snoeyink and Jenkins, 

1980). The potential obtained in ground waters is a mixed potential that reflects the potential of 
many reactions and cannot be used for quantitative interpretation (Stumm and Morgan, 1981). The 
approximate location of the contaminant plume can be identified in the field by measurement of the 
redox potential of the ground water. 

To overcome the limitations imposed by traditional redox measurements, recent work has 
focused on the measurement of molecular hydrogen to accurately describe the predominant in situ 
redox reactions (Chapelle et al., 1995; Lovley et al., 1994; Lovley and Goodwin, 1988). The evi¬ 
dence suggests that concentrations of H, in ground water can be correlated with specific microbial 
processes, and these concentrations can be used to identify zones of methanogenesis, sulfate reduc¬ 
tion, and iron reduction in the subsurface (Chapelle, 1996). 

Electron Acceptors: Measurement of the available electron acceptors is critical in identifying 
the predominant microbial and geochemical processes occurring in situ at the time of sample collec¬ 
tion. Nitrate and sulfate are found naturally in most ground waters and will subsequently be used as 
electron acceptors once oxygen is consumed. Oxidized forms of iron and manganese can be used as 
electron acceptors before sulfate reduction commences. Iron and manganese minerals solubilize 
coincidently with sulfate reduction, and their reduced forms scavenge oxygen to the extent that strict 
anaerobes (some sulfate reducers and all methanogens) can develop. Sulfate is found in many 
depositional environments, and sulfate reduction may be very common in many contaminated 
ground waters. In environments where sulfate is depleted, carbonate becomes the electron acceptor 
with methane gas produced as an end product. 

Temperature: The temperature at all monitoring wells should be measured to determine when 
the pumped water has stabilized and is ready for collection. Below approximately 30 feet, the 
temperature in the subsurface is fairly consistent on an annual basis. Microcosms should be stored at 
the average in situ temperature. Biological growth can occur over a wide range of temperatures, 
although most microorganisms are active primarily between 10°C and 35°C (50°F to 95°F). 

Chloride: Reductive dechlorination results in the accumulation of inorganic chloride. In 
aquifers with a low background of inorganic chloride, the concentration of inorganic chloride should 


C3-52 






increase as the chlorinated solvents are degraded. The sum of the inorganic chloride plus the chlo¬ 
ride in the contaminant being degraded should remain relatively consistent along the ground water 
flow path. 

Tables C.3.7 and C.3.8 list the geochemical parameters, contaminants, and daughter products 
that should be measured during site characterization for natural attenuation. The tables include the 
analyses that should be performed, the optimum range for natural attenuation of chlorinated solvents, 
and the interpretation of the value in relation to biological processes. 

Table C.3 .7 Geochemical Parameters Important to Microcosm Studies 


Analysis 

Range 

Interpretation 

Redox Potential 

<50 millivolt against 
Ag/AgCl 

Reductive pathway possible 

Sulfate 

<20 mg/L 

Competes at higher concentrations with reductive pathway 

Nitrate 

<1 mg/L 

Competes at higher concentrations with reductive pathway 

Oxygen 

<0.5 mg/L 

Tolerated, toxic to reductive pathway at higher concentrations 

Oxygen 

>1 mg/L 

Vinyl chloride oxidized 

Iron (II) 

>1 mg/L 

Reductive pathway possible 

Sulfide 

>1 mg/L 

Reductive pathway possible 

Hydrogen 

>1 nM 

Reductive pathway possible, vinyl chloride may accumulate 

Hydrogen 

<1 nM 

Vinyl chloride oxidized 

pH 

5 < pH < 9 

Tolerated range 


Table C.3.8 Contaminants and Daughter Products Important to Microcosm Studies 


Analysis 

Interpretation 

PCE 

Material spilled 

TCE 

Material spilled or daughter product of PCE 

1,1,1-TCA 

Material spilled 

cis- 1,2-DCE 

Daughter product of TCE 

trans-l ,2-DCE 

Daughter product of TCE 

Vinyl Chloride 

Daughter product of dichloroethylenes 

Ethene 

Daughter product of vinyl chloride 

Ethane 

Daughter product of ethene 

Methane 

Ultimate reductive daughter product 

Chloride 

Daughter product of organic chlorine 

Carbon Dioxide 

Ultimate oxidative daughter product 

Alkalinity 

Results from interaction of carbon dioxide with aquifer minerals 


C3-53 































C.3.4.6 Microcosm Construction 

During construction of the microcosms, it is best if all manipulations take place in an anaerobic 
glovebox. These gloveboxes exclude oxygen and provide an environment where the integrity of the 
core material may be maintained, since many strict anaerobic bacteria are sensitive to oxygen. Strin¬ 
gent aseptic precautions not necessary for microcosm construction. It is more important to maintain 
anaerobic conditions of the aquifer material and solutions added to the microcosm bottles. 

The microcosms should have approximately the same ratio of solids to water as the in situ 
aquifer material, with a minimum or negligible headspace. Most bacteria in the subsurface are 
attached to the aquifer solids. If a microcosm has an excess of water, and the contaminant is prima¬ 
rily in the dissolved phase, the bacteria must consume or transform a great deal more contaminant to 
produce the same relative change in the contaminant concentration. As a result, the kinetics of 
removal at field scale will be underestimated in the microcosms. 

A minimum of three replicate microcosms for both living and control treatments should be 
constructed for each sampling event. Microcosms sacrificed at each sampling interval are preferable 
to microcosms that are repetitively sampled. The compounds of regulatory interest should be added 
at concentrations representative of the higher concentrations found in the geochemical region of the 
plume being evaluated. The compounds should be added as a concentrated aqueous solution. If an 
aqueous solution is not feasible, dioxane or acetonitrile may be used as solvents. Avoid carriers that 
can be metabolized anaerobically, particularly alcohols. If possible, use ground water from the site 
to prepare dosing solutions and to restore water lost from the core barrel during sample collection. 

For long-term microcosm studies, autoclaving is the preferred method for sterilization. Nothing 
available to sterilize core samples works perfectly. Mercuric chloride is excellent for short-term 
studies (weeks or months). However, mercuric chloride complexes to clays, and control may be lost 
as it is sorbed over time. Sodium azide is effective in repressing metabolism of bacteria that have 
cytochromes, but is not effective on strict anaerobes. 

The microcosms should be incubated in the dark at the ambient temperature of the aquifer. It is 
preferable that the microcosms be incubated inverted in an anaerobic glovebox. Anaerobic jars are 
also available that maintain an oxygen-free environment for the microcosms. Dry redox indicator 
strips can be placed in the jars to assure that anoxic conditions are maintained. If no anaerobic 
storage is available, the inverted microcosms can be immersed in approximately two inches of water 
during incubation. Teflon®-lined butyl rubber septa are excellent for excluding oxygen and should 
be used if the microcosms must be stored outside an anaerobic environment. 

The studies should last from one year to eighteen months. The residence time of a plume may 
be several years to tens of years at field scale. Rates of transformation that are slow in terms of 
laboratory experimentation may have a considerable environmental significance. A microcosm 
study lasting only a few weeks to months may not have the resolution to detect slow changes that are 
of environmental significance. Additionally, microcosm studies often distinguish a pattern of se¬ 
quential biodegradation of the contaminants of interest and their daughter products. 

C.3.4.7 Microcosm Interpretation 

As a practical matter, batch microcosms with an optimal solids/water ratio, that are sampled 
every two months in triplicate, for up to eighteen months, can resolve biodegradation from abiotic 
losses with a detection limit of 0.001 to 0.0005 per day. Rates determined from replicated batch 
microcosms are found to more accurately duplicate field rates of natural attenuation than column 
studies. Many plumes show significant attenuation of contamination at field calibrated rates that are 
slower than the detection limit of microcosms. Although rate constants for modeling purposes are 
more appropriately acquired from field-scale studies, it is reassuring when the rates in the field and 
the rates in the laboratory agree. 


C3-54 


The rates measured in the microcosm study may be faster than the estimated field rate. This 
may not be due to an error in the laboratory study, particularly if estimation of the field-scale rate of 
attenuation did not account for regions of preferential flow in the aquifer. The regions of preferential 
flow may be determined by use of a downhole flow meter or by other methods for determining 
hydraulic conductivity in one- to two-foot sections of the aquifer. 

Statistical comparisons can determine if removals of contaminants of concern in the living 
treatments are significantly different from zero or significantly different from any sorption that is 
occurring. Comparisons are made on the first-order rate of removal, that is, the slope of a linear 
regression of the natural logarithm of the concentration remaining against time of incubation for both 
the living and control microcosm. These slopes (removal rates) are compared to determine if they are 
different, and if so, extent of difference that can be detected at a given level of confidence. 

C.3.4.8 The Tibbetts Road Case Study 

The Tibbetts Road Superfund Site in Barrington, New Hampshire, a former private home, was 
used to store drums of various chemicals from 1944 to 1984. The primary ground-water contami¬ 
nants in the overburden and bedrock aquifers were benzene and TCE, with respective concentrations 
of 7,800 pg/L and 1,100 pg/L. High concentrations of arsenic, chromium, nickel, and lead were also 
found. 

Material collected at the site was used to construct a microcosm study evaluating the removal of 
benzene, toluene, and TCE. This material was acquired from the most contaminated source at the 
site, the waste pile near the origin of Segment A (Figure C.3.10). Microcosms were incubated for 
nine months. The aquifer material was added to 20-mL headspace vials, dosed with 1 mL of spiking 
solution, capped with a Teflon®-lined, gray butyl rubber septa, and sealed with an aluminum crimp 
cap. Controls were prepared by autoclaving the material used to construct the microcosms over¬ 
night. Initial concentrations for benzene, toluene, and TCE were, respectively, 380 pg/L, 450 pg/L, 
and 330 pg/L. The microcosms were thoroughly mixed by vortexing, then stored inverted in the 
dark at the ambient temperature of 10°C. 

The results (Figures C.3.11, C.3.12, and C.3.13; Table C.3.9) show that significant biodegrada¬ 
tion of both petroleum aromatic hydrocarbons and the chlorinated solvent had occurred. Significant 
removal in the control microcosms also occurred for all compounds. The data exhibited more 
variability in the living microcosms than in the control treatment, which is a pattern that has been 
observed in other microcosm studies. The removals observed in the controls are probably due to 
sorption; however, this study exhibited more sorption than typically seen. 

The rate constants determined from the microcosm study for the three compounds are shown in 
Table C.3.10. The appropriate rate constant to be used in a model or a risk assessment would be the 
first-order removal in the living treatment minus the first-order removal in the control, in other words 
the removal that is in excess of the removal in the controls. 

The first-order removal in the living and control microcosms was estimated as the linear regres¬ 
sion of the natural logarithm of concentration remaining in each microcosm in each treatment against 
time of incubation. Student’s t-distribution with n-2 degrees of freedom was used to estimate the 
95% confidence interval. The standard error of the difference of the rates of removal in living and 
control microcosms was estimated as the square root of the sum of the squares of the standard errors 
of the living and control microcosms, with n-4 degrees of freedom (Glantz, 1992). 

Table C.3.11 presents the concentrations of organic compounds and their metabolic products in 
monitoring wells used to define line segments in the aquifer for estimation of field-scale rate con¬ 
stants. Wells in this aquifer showed little accumulation of trans- 1,2-DCE; 1,1-DCE; vinyl chloride; 
or ethene, although removals of TCE and cis- 1,2-DCE were extensive. This can be explained by the 
observation (Bradley and Chapelle, 1996) that iron-reducing bacteria can rapidly oxidize vinyl 
chloride to carbon dioxide. Filterable iron accumulated in ground water in this aquifer. 


C3-55 



Figure C.3.10 Tibbetts Road study site. 



□ TCE Microcosm 
■ TCE Control 


Figure C.3.11 TCE microcosm results. 


C3-56 

















□ Benzene Microcosm 
■ Benzene Conlrol 


1 -|-1-1-1-1-f-<-I-1-1 

0 5 10 15 20 25 30 35 40 45 

Time (Weeks) 


Figure C.3.12 Benzene microcosm results. 



□ Toluene Microcosm 
■ Toluene Conlrol 


C3-57 


















Table C.3.9 Concentrations (/ug/L) ofTCE, Benzene, and Toluene in the Tibbetts Road Microcosms 


Compound 

Time Zero 
Microcosm 

Time Zero 
Control 

Week 23 
Microcosm 

Week 23 
Control 

Week 42 
Microcosm 

Week 42 
Control 

TCE 

328 

337 

1 

180 

2 

36.3 


261 

394 

12.5 

116 

2 

54.5 


309 

367 

8.46 

99.9 

2 

42.3 

Mean ± 

Standard 

Deviation 

299 ± 34.5 

366 ± 28.5 

7.32 ± 5.83 

132 ± 42.4 

2.0 ± 0.0 

44.4 ± 9.27 








Benzene 

366 

396 

201 

236 

11.1 

146 


280 

462 

276 

180 

20.5 

105 


340 

433 

22.8 

152 

11.6 

139 

Mean ± 

Standard 

Deviation 

329 ±44.1 

430 ± 33.1 

167 ± 130 

189 ± 42.8 

14.4 ± 5.29 

130 ±21.9 








Toluene 

443 

460 

228 

254 

2 

136 


342 

557 

304 

185 

2.5 

92 


411 

502 

19.9 

157 

16.6 

115 

Mean ± 

Standard 

Deviation 

399 ±51.6 

506 ±48.6 

184 ± 147 

199 ±49.9 

7.03 ± 8.29 

114 ± 22.0 


The extent of attenuation from well to well listed in Table C.3.11, and the travel time between 
wells in a segment (Figure C.3.4) were used to calculate first-order rate constants for each segment 
(Table C.3.12). Travel time between monitoring wells was calculated from site-specific estimates of 
hydraulic conductivity and from the hydraulic gradient. In the area sampled for the microcosm study, 
the estimated Darcy flow was 2.0 feet per year. With an estimated porosity in this particular glacial 
till of 0.1, this corresponds to a plume velocity of 20 feet per year. 

C.3.4.9 Summary 

Table C.3.13 compares the first-order rate constants estimated from the microcosm studies to 
the rate constants estimated at field scale. The agreement between the independent estimates of rate 
is good; indicating that the rates can appropriately be used in a risk assessment. The rates of biodeg¬ 
radation documented in the microcosm study could easily account for the disappearance of trichloro¬ 
ethylene, benzene, and toluene observed at field scale. The rates estimated from the microcosm- 
study are several-fold higher than the rates estimated at field scale. This may reflect an underestima¬ 
tion of the true rate in the field. The estimates of plume velocity assumed that the aquifer was 
homogeneous. No attempt was made in this study to correct the estimate of plume velocity for the 
influence of preferential flow paths. Preferential flow paths with a higher hydraulic conductivity 
than average would result in a faster velocity of the plume, thus a lower residence time and faster 
rate of removal at field scale. 


C3-58 























Table C.3.10 First-order Rate Constants for Removal o/TCE, Benzene, and Toluene in the Tibbetts Road 
Microcosms 


Parameter 

Living Microcosms 

Autoclaved 

Controls 

Removal Above 

Controls 

First-order Rate of Removal (per year) 

TCE 

6.31 

2.62 

3.69 

95% Confidence Interval 

±2.50 

±0.50 

±2.31 

Minimum Rate Significant at 95% 

Confidence 



1.38 





Benzene 

3.87 

1.51 

2.36 

95% Confidence Interval 

±1.96 

±0.44 

±1.83 

Minimum Rate Significant at 95% 

Confidence 



0.53 





Toluene 

5.49 

1.86 

3.63 

95% Confidence Interval 

±2.87 

±0.45 

±2.64 

Minimum Rate Significant at 95% 

Confidence 

> 


0.99 


Table C.3.11 Concentrations of Contaminants and Metabolic By-products in Monitoring Wells along 
Segments in the Plume used to Estimate Field-scale Rate Constants 


Parameter 

Segment A 

Segment B 

Segment C 

Monitoring 

Well 

80S 

79S 

70S 

52S 

70S 

53S 


Upgradient Downgradient 

Up 

gradient 

Down 

gradient 

Upgradient 

Down 

gradient 


....(pg/liter)-- 

TCE 

200 

13.7 

710 

67 

710 

3.1 

cis- 1,2-DCE 

740 

10.9 

220 

270 

220 

2.9 

trans- 1,2-DCE 

0.41 

<1 

0.8 

0.3 

0.8 

<1 

1,1-DCE 

0.99 

<1 

<1 

1.6 

<1 

<1 

Vinyl Chloride 

<1 

<1 

<1 

<1 

<1 

<1 

Ethene 

<4 

<4 

7 

<4 

7 

<4 

Benzene 

510 

2.5 

493 

420 

493 

<1 

Toluene 

10000 

<1 

3850 

900 

3850 

<1 

o-Xylene 

1400 

8.4 

240 

71 

240 

<1 

w-Xylene 

2500 

<1 

360 

59 

360 

<1 

p-Xylene 

1400 

22 

1100 

320 

1100 

<1 

Ethylbenzene 

1300 

0.7 

760 

310 

760 

<1 

Methane 

353 

77 

8 

3 

8 

<2 

Iron 






27000 


C3-59 
























































Table C.3.12 First-order Rate Constants for Contaminant Attenuation in Segments of the Tibbetts Road 

Plume 


Flow Path Segments in Length and Time of Ground-water Travel 


Segment A 

130 feet = 6.5 years 

Segment B 

80 feet = 4.0 years 

Segment C 

200 feet = 10 years 

Compound 

First-order Rate Constants in Segments ( per year) 

TCE 

0.41 

0.59 

0.54 

cis- 1,2-DCE 

0.65 

produced 

0.43 

Benzene 

0.82 

0.04 

>0.62 

Toluene 

>1.42 

0.36 

>0.83 

o-Xylene 

0.79 

0.30 

>0.55 

w-Xylene 

>1.20 

0.45 

>0.59 

p-Xylene 

0.64 

0.31 

>0.70 

Ethylbenzene 

1.16 

0.22 

>0.66 


Table C.3.13 Comparison of First-order Rate Constants in a Microcosm Study, and in the Field, at the 
Tibbetts Road NPL Site 


Parameter 

Microcosms Corrected for 
Controls 

Field Scale 


Average 

Rate 

Minimum Rate 
Significant at 95% 
Confidence 

Segment A 

Segment B 

Segment C 


.First-order Rate (per year).. 

Trichloroethylene 

3.69 

1.38 

0.41 

0.59 

0.54 

Benzene 

2.36 

0.53 

0.82 

0.04 

>0.62 

Toluene 

3.63 

0.99 

>1.42 

0.36 

>0.83 


☆ 


U.S. GOVERNMENT PRINTING OFFICE: 1999 - 750-101/00014 


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